Sorption I: General Introduction and Sorption Processes ...mrh/ESE 176 SGI Book/SGI Chp 9 Sorption Part 1.pdfEnvironmental Organic Chemistry, 2nd Edition. Rene P. Schwarzenbach, Philip

2 75

Chapter 9

SORPTION I: GENERAL INTRODUCTION AND SORPTION PROCESSES INVOLVING ORGANIC MATTER

9.1 Introduction 9.2 Sorption Isotherms, Solid-Water Distribution Coefficients (Kid), and the

Fraction Dissolved K.,) Qualitative Considerations Quantitative Description of Sorption Isotherms The Solid-Water Distribution Coefficient Kid

Illustrative Example 9.1 : Determining Kid Values from Experimental Data Dissolved and Sorbed Fractions of a Compound in a System The Complex Nature of Kjd

9.3 Sorption of Neutral Organic Compounds from Water to Solid-Phase Organic Matter (POM) Overview Structural Characteristics of POM Relevant to Sorption Determination of Ki,, Values and Availability of Experimental Data Estimation of K,,, Values K,,, as a Function of Sorbate Concentration Illustrative Example 9.2: Evaluating the Concentration Dependence of

Illustrative Example 9.3 : Estimating Pore Water Concentrations in a

Effect of Temperature and Solution Composition on K,,, Illustrative Example 9.4: How Much Does the Presence of 20% Methanol in

Sorption of Phenanthrene to Soil and Sediment POM

Polluted Sediment

the “Aqueous” Phase Affect the Retardation of Phenanthrene in an Aqulfer?

Environmental Organic Chemistry, 2nd Edition. Rene P. Schwarzenbach, Philip M. Gschwend and Dieter M. Imboden

Copyright 0 2003 John Wiley &L Sons, Inc.

276 Sorption I: Sorption Processes Involving Organic Matter

9.4 Sorption of Neutral Compounds to “Dissolved” Organic Matter (DOM) Qualitative Description of DOM-Solute Associations Determination of KjDoc Values and Availability of Experimental Data DOM Properties Governing the Magnitude of KjDoc Effect of pH, Ionic Strength, and Temperature on KiDoc LFERs Relating KiDoc Values to K,, Values Illustrative Example 9.5: Evaluating the Effect of DOMon the

Bioavailability of Benzo(a)pyrene

9.5 Sorption of Organic Acids and Bases to Natural Organic Matter (NOM) Effect of Charged Moieties on Sorption: General Considerations Sorption of Compounds Forming Anionic Species (Organic Acids) Sorption of Compounds Forming Cationic Species (Organic Bases)

9.6 Questions and Problems

Introduction 2 77

Introduction

The process in which chemicals become associated with solid phases is generally referred to as sorption. It is adsorption if the molecules attach to a two-dimensional surface, while it is absorption if the molecules penetrate into a three-dimensional matrix. This phase transfer process may involve vapor molecules or dissolved molecules associating with solid phases.

Sorption is extremely important because it may dramatically affect the fate and impacts of chemicals in the environment. Such importance is readily understood if we recognize that structurally identical molecules behave very differently if they are: (a) in the gas phase or (b) surrounded by water molecules and ions as opposed to (c) clinging onto the exterior of solids or (d) buried within a solid matrix (Fig. 9.1). Clearly, the environmental transport of waterborne molecules must differ from the movements of the same kind of molecules attached to particles that settle. Also, transport of a given compound in porous media such as soils, sediments, and aquifers is strongly influenced by the compound’s tendency to sorb to the various components of the solid matrix. Additionally, only dissolved molecules are available to collide with the interfaces leading to other environmental compartments such as the atmosphere; thus phase transfers are controlled by the dissolved species of a chemical (Chapters 19 and 20). Similarly, since molecular transfer is a prerequisite for the uptake of organic pollutants by organisms, the bioavailability of a given compound and thus its rate of biotransformation or its toxic effect(s) are affected by sorption processes (Chapters 10 and 17). Furthermore, some sorbed molecules are substantially shaded from incident light; therefore, these molecules may not experience direct photolysis processes. Moreover, when present inside solid matrices, they may never come in contact with short-lived, solution-phase photooxidants like OH-radicals (Chapters 15 and 16). Finally, since the chemical natures of aqueous solutions and solid environments differ greatly (e.g., pH, redox conditions), various chemical reactions including hydrolysis or redox reactions may occur at very different rates in the sorbed and dissolved states (Chapters 13 and 14). Hence, we must understand solid-solution and solid-gas phase exchange phenomena before we can quantify virtually any other process affecting the fates of organic chemicals in the environment.

Unfortunately, when we are dealing with natural environments, sorption is very often not an exchange between one homogeneous solution/vapor phase and a single solid medium. Rather, in a given system some combination of interactions may govern the association of a particular chemical (called the sorbate) with any particular solid or mixture of solids (called the soubent(s)). Consider the case of 3,4-dimethylaniline (3,4-dimethyl aminobenzene, Fig. 9.2). This compound is a weak base with pKi, = 5.28 (see Illustrative Example 8.2); hence, it reacts in aqueous solution to form some 3,4-dimethyl ammonium cations. For the fraction of molecules that remain uncharged, this organic compound may escape the water by penetrating the natural organic matter present in the system. Additionally, such a nonionic molecule may displace water molecules from the region near a mineral surface to some extent and be held there by London dispersive and polar interactions. These two types of sorption mechanisms are general and will operate


t

dissolved organic molecules colliding with air-water interface

and volatilizing

settling with particles for sorbed organic molecules

air

ti H H

water

Figure 9.1 Illustration of some processes in which sorbed species behave differently from dissolved molecules of the same substance. (a) Dissolved species may participate directly in air-water exchange while sorbed species may settle

may react at different rates as com-

parts due to differential access of other dissolved and solid-phase "reactants."

.O - H

H 0 - H H H.O Y .O-H .O ~ - 6 ~ H:o H H*O.

H H .O'H 3 H H*O with solids. (b) Dissolved species 9 ,H H.o

pared with their sorbed counter- H H .H ti H H'O J.1 H b

do H - 0

H'O'H H Q b H H -0

H

dissolved organic molecules are more accessible to light, to other dissolved chemicals, and to microorganisms than sorbed molecules

for any organic chemical and any natural solid. Additionally, since the sorbate is ionizable in the aqueous solution, then electrostatic attraction to specific surface sites exhibiting the opposite charge will promote sorption of the ionic species. Finally, should the sorbate and the sorbent exhibit mutually reactive moieties (e.g., in Fig. 9.2 a carbonyl group on the sorbent and an amino group on the sorbate), some portion of a chemical may actually become bonded to the solid. All of these interaction mechanisms operate simultaneously, and the combination that dominates the overall solution-solid distribution will depend on the structural properties of the organic sorbate and the solid sorbent of interest.

In this and the following two chapters, we will focus on solid-aqueous solution and solid-air exchange involving natural sorbents. We will try to visualize the sets of molecular interactions involved in each of the above-mentioned sorption processes. With such pictures in our minds, we will seek to rationalize what makes various sorption mechanisms important under various circumstances. Establishing the critical compound properties and solid characteristics will enable us to understand

Introduction 279

Figure 9.2 Some sorbent-sorbate interactions possibly controlling the association of a chemical, (3,4- dimethylaniline), with natural solids.

when and to what extent predictive approaches for quantification of sorption may be applied. Ultimately, we should gain some feeling for what structural features of a chemical and what characteristics of solids (and solutions) are important to sorptive interactions. In this chapter, we will focus on sorption processes involving natural organic matter. In Chapter 10 we will address sorption to living media, that is, we will treat bioaccumulation. Finally, sorption from water to mineral surfaces and gas-solid phase transfers will be discussed in Chapter 11. To start out, we will first address some general aspects needed to quantify sorption equilibrium in a given system. Kinetic aspects of sorption processes will be treated in Chapter 19.


Figure 9.3 Various types of observed relationships between concentrations of a chemical in the sorbed state, C,,, and the dissolved state, C,w. Note that similar relationships apply to the sorption of gaseous compounds to solid sorbents.

Sorption Isotherms, Solid-Water Distribution Coefficients (Kid), and the Fraction Dissolved &,)

Qualitative Considerations

When we are interested in the equilibrium distribution of a chemical between the solids and solution present in any particular volume of an aquatic environment, we begin by considering how the total sorbate concentration, Cis (e.g., mol . kg-I), depends on chemical’s concentration in the solution, Ci, (e.g., mol .L-’). The relationship of these two concentrations is commonly referred to as a sorption isotherm. The name isotherm is used to indicate that this sorption relationship applies only at a constant temperature.

Experimentally determined sorption isotherms exhibit a variety of shapes for diverse combinations of sorbates and sorbents (Fig. 9.3). The simplest case (Fig. 9 . 3 ~ ) is the one in which the afinity of the sorbate for the sorbent remains the same over the observed concentration range. This is the so-called linear isotherm case. It applies to situations where partitioning into a homogeneous organic phase is dominating the overall sorption, and/or at low concentrations where the strongest adsorption sites are far from being saturated. The second types of behavior (Figs. 9.3b and c) reflect those situations in which at higher and higher sorbate concentrations it becomes more and more difficult to sorb additional molecules. This occurs in cases where the binding sites become filled and/or remaining sites are less attractive to the sorbate molecules. In the extreme case (Fig. 9 . 3 ~ ) ~ above some maximum Cis value, all sites are “saturated” and no more additional sorption is possible. Isotherms of the type shown in Figs. 9.3b and c are encountered in studies of adsorption processes to organic (e.g., activated carbon) or inorganic (e.g., clay mineral) surfaces. Of course, in a soil or sediment, there may be more than one important sorbent present. Therefore, the overall sorption isotherm may reflect the superposition of several individual isotherms that are characteristic for each specific type of sorbent. When such a case involves an adsorbent (e.g., soot, clay mineral) exhibiting a limited number of sites with a high affinity for the sorbate (type (c) isotherm) that dominates the overall sorption at low concentrations, plus a partitioning process (e.g., into natural organic matter; type (a) isotherm) predominating at higher concentrations, then a mixed isotherm is seen (Fig. 9.3b or 6). Likewise, superimposition of multiple adsorption isotherms results in a mixed isotherm looking like an isotherm of type (b) (Weber et al., 1992).

Another case that is less frequently encountered involves the situation in which previously sorbed molecules lead to a modification of the sorbent which favors further sorption (Fig. 9.3e). Such effects have been seen in studies involving anionic or cationic surfactants as sorbates. In some of these cases, a sigmoidal isotherm shape (Fig. 9.3f) has been observed, indicating that the sorption-promoting effect starts only after a certain loading of the sorbent.

In summary, depending on the composition of a natural bulk sorbent and on the chemical nature of the sorbate, multiple sorption mechanisms can act simultaneously and the resulting isotherms may have a variety of different shapes. We should note that it is not possible to prove a particular sorption mechanism

Sorption Isotherms 281

applies from the shape of the isotherm. Nevertheless, the isotherm type and its degree of nonlinearity must be consistent with the sorption mechanism(s) prevailing in a given situation.

Quantitative Description of Sorption Isotherms

A very common mathematical approach for fitting experimentally determined sorption data using a minimum of adjustable parameters employs an empirical relationship known as the Freundlich isotherm:

where KIF is the Freundlich constant or capacity factor [(e.g., Eq. 9-1 in (mol. kg-') (mol . L-')-nf)]; and n, is the Freundlich exponent. Note that for a correct thermo- dynamic treatment of Eq. 9-1 we would always have to use dimensionless activities of compound i in both the sorbed and aqueous phase in order to obtain a dimensionless Kz. However, in practice C,, and Ciw are expressed in a variety of concentration units. Therefore, KIF is commonly reported in the corresponding units, which also means that for n, f 1, KIF depends nonlinearly on the units in which C,, is expressed (see Illustrative Example 9.1 and Problem 9.5).

The relationship Eq. 9-1 assumes there are multiple types of sorption sites acting in parallel, with each site type exhibiting a different sorption free energy and total site abundance. The exponent is an index of the diversity of free energies associated with the sorption of the solute by multiple components of a heterogeneous sorbent (Weber and Digiano, 1996). When n, = 1, the isotherm is linear and we infer constant sorption free energies at all sorbate concentrations (Fig. 9.3a); when ni < 1, the isotherm is concave downward and one infers that added sorbates are bound with

' weaker and weaker free energies (Fig. 9.3b); finally when n, > 1, the isotherm is convex upward and we infer that more sorbate presence in the sorbent enhances the

experimental data by linear regression of the logarithmic form of Eq. 9-1 (Fig. 9.4; see also Illustrative Example 9.1):

I free energies of firther sorption (Fig. 9.3e). KiF and n, can be deduced from 0

log 4,

Figure 9.4 Graphic representation of the Freundlich isotherm Eq. 9-2 for the three cases n, > 1, n, = 1, and

(9-2)

< '. Note that nz and log KiF are If a given isotherm cannot be described by Eq. 9-2, then some assumptions behind obtained from the slope (n,) and intercept (log K , ~ indicated by the the Freundlich multi-site conceptualization are not valid. For example, if there are points at log C,, = 0) of the reg- limited total sorption sites that become saturated (case shown in Fig. 9 . 3 ~ ) ~ then C,, ression line. cannot increase indefinitely with increasing C,,. In this case, the Langmuir isotherm

may be a more appropriate model:

where rmax represents the total number of surface sites per mass of sorbent. In the ideal case, rmax would be equal for all sorbates. However, in reality, rmax may vary somewhat between different compounds (e.g., because of differences in sorbate size). Therefore, it usually represents the maximum achievable surface concentration of a given compound i (i.e., rmax = C,,,,,). The constant KiL, which is


commonly referred to as the Langmuir constant, is defined as the equilibrium constant of the sorption reaction:

surface site + sorbate in aqueous solution & sorbed sorbate

Note that in this approach, since KiL is constant, this implies a constant sorbate affinity for all surface sites. To derive KiL and Cis,,,, from experimental data, one may fit

.C,,LX l/Ciw versus l/Cis:

Figure 9.5 Graphic representation of the Lanrrmuir isotherm Ea. 9-4.

1 1 (9-4)

., and use the slope and intercept to extract estimates of the isotherm constants (Fig. 9.5).

There are many cases in which the relationship between sorbed concentrations and dissolved concentrations covering a large concentration range cannot be described solely by a linear, a Langmuir, or even a Freundlich equation (e.g., cases d andfin Fig. 9.3). In these cases, combinations of linear-, Langmuir-, and/or Freundlich-type equations may need to be applied (e.g., Weber et al., 1992; Xing and Pignatello, 1997; Xia and Ball, 1999). Among these distributed reactivity models (Weber et al., 1992), the simplest case involves a pair of sorption mechanisms involving absorption (e.g., linear isotherm with partition coefficient, Kip) and site-limited adsorption (e.g., Langmuir isotherm), and the resultant combined equation is:

Note that Cis,,,, and K,L can be derived from the slope and intercept of the regression line (see also Illustrative Example 9.1).

Another form that fits data from sediments known to contain black carbon (e.g., soot) uses a combination of a linear isotherm and a Freundlich isotherm (Accardi- Dey and Gschwend, 2002):

C i s = K j p C i w + K~FC: (9-6)

These dual-mode models have been found to be quite good in fitting experimental data for natural sorbents that contain components exhibiting a limited number of more highly active adsorption sites as well as components into which organic compounds may absorb (Huang et al., 1997; Xing and Pignatello, 1997; Xia and Ball, 1999). At low concentrations, the Langmuir or the Freundlich term may dominate the overall isotherm, while at high concentrations (e.g., KiL . Ciw n l), the absorption term dominates (see Section 9.3).

The Solid-Water Distribution Coefficient, Kid

To assess the extent to which a compound is associated with solid phases in a given system at equilibrium (see below), we need to know the ratio of the compound's total equilibrium concentrations in the solids and in the aqueous solution. We denote this solid-water distribution coefficient as Kid (e.g., in L. kg-' solid):

Lis K . -- id - c i w

(9-7)


(When writing natural solid-water distribution or partition coefficients, we will use a somewhat different subscript terminology than used for air-water or organic solvent-water partitioning; that is, we will not indicate the involvement of a water phase by using a subscript "w".) When dealing with nonlinear isotherms, the value of this ratio may apply only at the given solute concentration (i.e., if n, in Eq. 9-1 is substantially different from 1). Inserting Eq. 9-1 into Eq. 9-7, we can see how Kid varies with sorbate concentration:

For practical applications, one often assumes that Kid is constant over some concentration range. We can examine the reasonableness of such a simplification by differ- entiating Kid with respect to Ciw in Eq. 9-8 and rearranging the result to find:

So the assumption about the constancy Of Kid is equivalent to presuming either: (a) the overall process is described by a linear isotherm (n, - 1= 0), or (b) the relative concentration variation, (dCiw/Ciw), is sufficiently small that when multiplied by (n, - 1) the relative Kid variation, (dKjd/Kid), is also small. For example, if the sorbate concentration range is less than a factor of 10, when multiplied by (n, - 1) with an n, value of 0.7, then the solid-water distribution coefficient would vary by less than a factor of 3 .

Illustrative Example 9.1

NO,

1,4-dinitrobenzene (1,4-DNB)

0.06 97 0.17 24 1 0.24 363 0.34 483 0.51 633 0.85 915 1.8 1640 2.8 2160 3.6 2850 7.6 4240

19.5 6100 26.5 7060

Determining Kid Values from Experimental Data

A common way to determine Kid values is to measure sorption isotherms in butch experiments. To this end, the equilibrium concentrations of a given compound in the solid phase (Cis) and in the aqueous phase (C,J are determined at various compound concentrations and/or solid-water ratios. Consider now the sorption of 1,4- dinitrobenzene (1,4-DNB) to the homoionic clay mineral, K'-illite, at pH 7.0 and 20°C. 1,4-DNB forms electron donor-acceptor (EDA) complexes with clay minerals (see Chapter 11). In a series of batch experiments, Haderlein et al. (1996) measured the data at 20°C given in the margin.

Problem

Using this data, estimate the K,,-values for 1,4-DNB in a K'-illite-water suspension (pH 7.0 at 20°C) for equilibrium concentrations of 1,4-DNB in the aqueous phase of 0.20 pM and of 15 pM, respectively.

Answer

Plot Cis versus Cjw to see the shape of the sorption isotherm (Fig. 1):

For Kid at Ciw = 0.20 pM, assume a h e a r isotherm for the concentration range 0-0.5 pM. Perform a least squares fit of Cis versus C, using only the first four data


Figure 1 Plot of Cis versus C,. The dotted line represents the fitted Langmuir equation (see below).

Figure 2 Plot of log C,, versus log C,,using the whole data set.

8000 I

C 6000 b, Y

. ci” 2000 1 / + *

I , ’ 0 0 L 10

200 ~T2Tl 100 0 0 0.1 0.2 0.3

20 30

1.51 I I I I I

-1.5 -1 -0.5 0 0.5 1 1.5

log Ciwl pM

points and the origin (see insert in Fig. 1). The resulting regression equation is:

Cis = 1425 Ciw (R’ = 1.0)

Hence, you get a I ( d value (slope) of 1425 L.kg-’ that is valid for the whole concentration range considered (i.e., Ciw I 5 pM).

For deriving Kid at C, = 15 pM, fit the experimental data with the Freundlich equation (Eq. 9.1). To determine the KiF and n, values use Eq. 9.2 (i.e., perform a least squares fit of log Cis versus log C, using all data points).

The resulting regression here is:

log Cis = 0.70 log Cjw + 2.97 (R’ = 0.98)


Figure 3 Plot of UCis versus l/C, for the data points with C, > 0.5 pM.

0.001 2

0.001 0

- h

I

,- 0.0008 b, Y

0 - 5 0.0006

2 0.0004

v -. ul

0.0002

0

0 0.2 0.4 0.6 0.8 1 1.2

l/Ciw / ( pM -')

Hence, KiF = 102.97 2 1000 (pmol.kg-' ,uM-'.~') [see comment on units of KiF below Eq. 9-11 and ni = 0.70; therefore (Eq. 9-8):

Kid = 1000. c,;0.3 Insertion of Ciw = 15 pM yields a value of about 450 L . kg-'.

Note that this K;d value is significantly smaller than the Kid obtained in the linear part of the isotherm (i.e., at low 1,4-DNB concentrations). Furthermore, as can be seen from Fig. 2, the Freundlich equation overestimates Cis (and thus Kid) at both the low and the high end of the concentration range considered. In fact, inspection of Fig. 2 reveals that at very high concentrations, the IS'-illite surface seems to become saturated with 1,4-DNB, which is not surprising considering that only limited adsorption sites are available. In such a case, the sorption isotherm can also be approximated by a Langmuir equation (Eq. 9-3).

To get the corresponding KiL and Cimaxvalues, use Eq. 9-4 (i.e., perform a least squares fit of l/Cis versus l/Cjw). Use only the data with Ciw > 0.5 pM to get a reasonable weighting of data points in the low and high concentration range.

The resulting regression equation is:

1 1 - = 0.000753-+0.000152 ( R 2 = 0.99) Cis C i w

yielding a C,,,, of 6600 pmol. kg-' and a KiL value of 0.201 L .,umol-'. At very low


concentrations (i.e., KiL. Ciw (( I), which includes C, = 0.20 pM, Kid is given by the linear relationship:

K;d=KiL'C;max= (0.201) (6600)= 1.?20L.kg-'

which is somewhat smaller than the K;d value determined from the linear regression analysis using only the first four data points (i.e., Kid = 1425 L. kg-', see above). This is not too surprising when considering that the Langmuir model assumes that all surface sites exhibit the same affinities for the sorbate. This is not necessarily the case, as it is likely that sites with higher affinities are occupied first. Therefore, a linear fit of data points determined at low concentrations can be expected to yield a higher apparent sorption coefficient as compared to the coeMicient calculated from nonlinear extrapolation of data covering a wide concentration rate.

Inserting C, = 15 ,uM into Eq. 9-3 with the above derived KiL and C,,, values yields a Ci,valueof(6600)(0.201)(15)/[1+(0.201)(15)] =4950pmol.kg-', andthusaKidof 4950 / 15 = 330 L . kg-'. This value is somewhat smaller than the one derived from the Freundlich equation (450 L . kg-'; see above). These calculations show that when estimating Kid values from experimental data, depending on the concentration range of interest, one has to make an optimal choice with respect to the selection of the experimental data points as well as with respect to the type of isotherm used to fit the data.

Dissolved and Sorbed Fractions of a Compound in a System

Armed with a Kid for a case of interest, we may evaluate what fraction of the compound is dissolved in the water,f;,, for any environmental volume containing both solids and water, but only these phases:

Ciw . Vw Jw = ClWV, + C1,M,

(9-10)

where Vw is the volume of water (e.g., L) in the total volume vat, and M, is the mass of solids (e.g., kg) present in that same total volume. Now if we substitute the product Kid. C, from Eq. 9-7 for Cis in Eq. 9- 10, we have:

C l W v w

Jw = CiwVw + KldCiwMs

- v w - v w -I- KidMs

(9-11)

Finally, noting that we refer to the quotient, Ms/Vw, as the solid-water phase ratio, r,, (e.g., kg .L~-') in the environmental compartment of interest, we may describe the fraction of chemical in solution as a fimction of Kid and this ratio:


(9-12)

Such an expression clearly indicates that for substances exhibiting a great affinity for solids (hence a large value of Kid) or in situations having large amounts of solids per volume of water (large value of r,,), we predict that correspondingly small fractions of the chemical remain dissolved in the water. Note the fraction associated with solids, As, must be given by ( 1-AW) since we assume that no other phases are present (e.g., air, other immiscible liquids).

The fraction of the total volume, Go, , that is not occupied by solids, theporosity, 4, is often used instead of rsw to characterize the solid-water phase ratio in some environmental systems like sediment beds or aquifers. In the absence of any gas phase, 9 is related to parameters discussed above by:

(9-13)

where, V, , the volume occupied by particles, can be calculated fromMs/ps (where ps is the density of the solids and is typically near 2.5 kg L-' for many natural minerals.) Thus, we find the porosity is also given by:

1 - - v w 9 = v w +Ms/p, l+&wlps

and solving for r,, yields the corresponding relation:

1-4 9

r,w = ps-

(9-14)

(9-15)

Finally, in the soil and groundwater literature, it is also common to use still a third parameter called bulk density, A. Bulk density reflects the ratio, M,/V,, , so we see it is simply given by p, (1- 4). Thus, knowing bulk density we have r,, is equal to A/#. It is a matter of convenience whether r,,, 4 , or pb is used.

The application of such solution- versus solid-associated speciation information may be illustrated by considering an organic chemical, say 1,4-dimethylbenzene (DMB), in a lake and in flowing groundwater. In lakes, the solid-water ratio is given by the suspended solids concentration (since Vw = yet), which is typically near lo4 kg . L-'. From experience we may know that the Kid value for DMB in this case happens to be 1 L . kg-'; therefore we can see that virtually all of this compound is in the dissolved form in the lake:

In contrast, now consider the groundwater situation; ps for aquifer solids is about 2.5 kg .L-' (e.g., quartz density is 2.65 kg.L-'); @is often between 0.2 and 0.4. If in our


Figure 9.6 Illustration of the retardation of 1,4-dimethylbenzene (DMB) transport in groundwater due to: (1) reversible sorptive exchange between water and solids, and ( 2 ) limiting transport of DMB to that fraction remaining in the flowing water. As dissolved molecules move ahead, they become sorbed and stopped, while molecules sorbed at the rear return to the water and catch up. Thus, overall transport of DMB is slower than that of the water itself.

particular groundwater situation $J = 0.2, and Y,, = 10 kg.L-', we predict that the fraction of DMB in solution, again assuming Kid of 1 L. kg-', is drastically lower than in the lake:

1 0.09 1

1+10.1 &,=-----

So we deduce that only one DMB molecule out of 11 will be in the moving groundwater at any instant (Fig. 9.6). This result has implications for the fate of the DMB in that subsurface environment. If DMB sorptive exchange between the aquifer solids and the water is fast relative to the groundwater flow and if sorption is reversible, we can conclude that the whole population of DMB molecules moves at one-eleventh the rate of the water. The phenomenon of diminished chemical transport speed relative to the water seepage velocity is referred to as retardation. It is commonly discussed using the retardationfactor, Rf, which is simply equal to the reciprocal of the fraction of molecules capable of moving with the flow at any instant, A;' (see Chapter 25).

Many situations require us to know something about the distribution of a chemical between a solution and solids. Our task then is to see how we can get Kid values suited for the cases that concern us. As we already pointed out above, these Kjd values are determined by the structures of the sorbates as well as the composition of the aqueous phase and the sorbents.


The Complex Nature of Kid The prediction of Kid for any particular combination of organic chemical and solids in the environment can be diEcult, but fortunately many situations appear reducible to fairly simple limiting cases. We begin by emphasizing that the way we defined Kid means that we may have lumped together many chemical species in each phase. For example, referring again to Fig. 9.2, we recognize that the total concentration of the dimethylaniline in the sorbed phase combines the contributions of molecules in many different sorbed forms. Even the solution in this case contains both a neutral and a charged species of this chemical. Thus, in a conceptual way, the distribution ratio for this case would have to be written as:

where C,,, is the concentration of sorbate i associated with the natural organic matter (expressed as organic carbon) present (mol . kg-' oc)

f,, is the weight fraction of solid which is natural organic matter (expressed as organic carbon, i.e., kg oc kg-' solid)

Cimin is the concentration of sorbate i associated with the mineral surface (mol . m-2)

Asurf is the specific surface area of the relevant solid

CieX is the concentration of ionized sorbate drawn towards positions of opposite charge on the solid surface (mol . mol-' surface charges)

oSudex is the net concentration of suitably charged sites on the solid surface (mol surface charges. m-2) for ion exchange

C,, is the concentration of sorbate i bonded in a reversible reaction to the solid (mol . mol-' reaction sites)

osurfrxn is the concentration of reactive sites on the solid surface (mol reaction sites. m")

C i , neut is the concentration of uncharged chemical i in solution (mol . L-')

C i , ion is the concentration of the charged chemical i in solution (mol . L-')

All terms in Eq. 9-16 may also deserve further subdivision. For example, C,,,.f,, may reflect the sum of adsorption and absorption mechanisms acting to associate the chemical to a variety of different forms of organic matter (e.g., living biomass of microorganisms, partially degraded organic matter from plants, plastic debris from humans, etc.). Similarly, Cimin.Asurf may reflect a linear combination of the interactions of several mineral surfaces present in a particular soil or sediment with a single sorbate. Thus, a soil consisting of montmorillonite, kaolinite, iron oxide,


and quartz mineral components may actually have Cimin .A,, = C,,,, . a .Asurf + Cikao. b .Aswf + Ciiron o x . c .Asurf + C,,,, . d.Asurf where the parameters a, b, c, and d are the area fractions exhibited by each mineral type. Similarly, Ci,, . om,. A,, may reflect bonding to several different kinds of surface moieties, each with its own reactivity with the sorbate (e.g., 3,4-dimethylaniline). For now, we will work from the simplified expression which is Eq. 9-16, primarily because there are few data available allowing rational subdivisions of soil or sediment differentially sorbing organic chemicals beyond that reflected in this equation.

It is very important to realize that only particular combinations of species in the numerator and denominator of complex Kid expressions like that of Eq. 9-16 are involved in any one exchange process. For example, in the case of dimethylaniline (DMA) (Fig. 9.2), exchanges between the solution and the solid-phase organic matter:

(9-17)

reflect establishing the same chemical potential of the uncharged DMA species in the water and in the particulate natural organic phase. As a result, a single free energy change and associated equilibrium constant applies to the sorption reaction de- picted by Eq. 9-17. Similarly, the combination:

(9-18)

would indicate a simultaneously occurring exchange of uncharged aniline molecules from aqueous solution to the available mineral surfaces. Again, this exchange is characterized by a unique free energy difference reflecting the equilibria shown in Eq. 9- 18. Likewise, the exchange of:

(9-19)

should be considered if it is the neutral sorbate which can react with components of the solid. Note that such specific binding to a particular solid phase moiety may prevent rapid desorption, and therefore such sorbate-solid associations may cause part or all of the sorption process to appear irreversible on some time scale of interest.

So far we have considered sorptive interactions in which only the DMA species was directly involved. In contrast, it is the charged DMA species (i.e., anilinium ions) that is important in the ion exchange process:

Sorption from Water to Solid-Phase Organic Matter (POM) 291

(9-20)

Of course, the anilinium ion in solution is quantitatively related to the neutral aniline species via an acid-base reaction having its own equilibrium constant (see Chapter 8). But we also emphasize that the solution-solid exchange shown in Eq. 19-20 has to be described using the appropriate equilibrium expression relating corresponding species in each phase. The influence of each sorption mechanism is ultimately reflected by all these equilibria in the overall expression, and each is weighted by the availability of the respective sorbent properties in the heterogeneous solid (i.e., A,, o,, om, or the various Asurf values). By combining information on the individual equilibria (e.g., Eqs. 9-17 through 9-20) with these sorbent properties, we can develop versions of the complex Kid expression (Eq. 9-1 6) which take into account the structure of the chemical we are considering. In the following, we discuss these individual equilibrium relationships.

Sorption of Neutral Organic Compounds from Water to Solid-Phase Organic Matter (POM)

Overview

Among the sorbents present in the environment, organic matter plays an important role in the overall sorption of many organic chemicals. This is true even for compounds that may undergo specific interactions with inorganic sorbent components (see Chapter 11). We can rationalize this importance by recognizing that most surfaces of inorganic sorbents are polar and expose a combination of hydroxy- and oxy-moieties to their exterior. These polar surfaces are especially attractive to substances like water that form hydrogen bonds. Hence, in contrast to air-solid surface partitioning (Section 11.2), the adsorption of a nonionic organic molecule from water to an inorganic surface requires displacing the water molecules at such a surface. This is quite unfavorable from an energetic point of view. However, absorption of organic chemicals into natural organic matter or adsorption to a hydrophobic organic surface does not require displacement of tightly bound water molecules. Hence, nonionic organic sorbates successfully compete for associations with solid-phase organic matter.

Therefore, we may not be too surprised to find that nonionic chemicals show increasing solid-water distribution ratios for soils and sediments with increasing amounts of natural organic matter. This is illustrated for tetrachloromethane (carbon tetrachloride, CT) and 1,2-dichlorobenzene (DCB) when these two sorbates were examined for their solid-water distribution coefficients using a large number of soils and sediments (Fig. 9.7, Kile et al., 1995.)

Note that the common analytical methods for determining the total organic material present in a sorbent often involve combusting the sample and measuring evolved


25

20 h 7

b

5 1 5 n k-

10

Figure 9.7 Observed increase in solid-water distribution ratios for the apolar compounds, tetrachloromethane (0) and 1,2-dichlorobenzene (A) with increasing organic matter content of the solids (measured as organic carbon, hc, see Eq. 9-21) for 32 soils and 36 sediments. Data from Kile et al. (1 995).

-

-

-

-

CI I

CI- c- CI I ct

tetrachloromethane

A A

A

A

A

A

A

A A A A

5 0

0 0 0.01 0.02 0.03 0.04 0.05 0.06 0.07

foc

COz. Therefore, the abundance of organic material present is often expressed by the weight fraction that consisted of reduced carbon:

mass of organic carbon total mass of sorbent

foc = (kg oc . kg-' solid) (9-21)

Obviously, it is actually the total organic mass consisting of carbon, hydrogen, oxygen, nitrogen, etc. within the solid phase that acts to sorb the chemical of interest (i.e., theh,,, in kg 0m.kg-l solid). Natural organic matter is typically made up of about half carbon (40 to 60% carbon); hence, fOm approximately equals 2 .f, and these two metrics are reasonably correlated.

Returning to the sorption observations (Fig. 9.7), as the mass fraction of organic carbon&,, present in the solids approaches zero, the Kid values for both compounds become very small. Even at very lowf,, values (i.e.,f,, 0.001 kg oc.kg-' solid), sorption to the organic components of a natural sorbent may still be the dominant mechanism (see Chapter 11).

In order to evaluate the ability of natural organic materials to sorb organic pollutants, it is useful to define an organic carbon normalized sorption coefficient:

(9-22)

where Cioc is the concentration of the total sorbate concentration associated with the natural organic carbon (i.e., mol-kg-' oc). Note that in this case, it is assumed that organic matter is the dominant sorbent; that is, Cis is given by C,,, . f,,, the first term in the numerator of Eq. 9-16. Clearly the value ofKjoc differs for tetrachloromethane and 172-dichlorobenzene (the slopes differ in Fig. 9.7), and it is generally true that


CT DCB

log Kioc / (L - kg-loc) log Kioc/ (L - kg-‘oc)

Figure 9.8 Frequency diagrams showing the variability in the log Kioc values of (a) tetrachloromethane (CT) and (b) 1,2-dichlorobenzene (DCB) for 32 soils (dark bars) and 36 sediments (light bars). The range ofLC values of the soils and sediments investigated is indicated in Fig. 9.7. Data from Kile et al. ( I 995).

each chemical has its own “organic carbon normalized” solid-water partition coefficient, Kim.

The Kid value of a given compound shows some variation between different soils and sediments exhibiting the same organic carbon content (Fig. 9.7). This indicates that not only the quantity, but also the quality of the organic material present has an influence on Kid. Normalizing to the organic carbon contents of each soil and sediment, we can examine this variability for both tetrachloromethane and 1,2- dichlorobenzene sorbing to a variety of soils and sediments of very different origins (Fig. 9.8.) All the Kim values lie within a factor of about 2 (i.e., f 2 0 - f 0.3 log units). We should emphasize that these data include only Kim values determined in the linear range of the isotherms by a single research group. The data show that for these two apolar compounds, soil organic matter seems on average to be a somewhat poorer “solvent” as compared to sediment organic matter (Fig. 9.8). In fact, the average Kcroc values are 60 f 7 L . kg-‘ oc for the 32 soils and 100 f 11 L . kg-’ oc for the 36 sediments investigated; similarly the average KDCBoc values are 290 & 42 L.kg-’ oc and 500 & 66 L.kg-’ oc, for the soils and sediments, respectively. Apparently, the sources of organic matter in terrestrial settings leave residues that are somewhat more polar than the corresponding residues derived chiefly in water bodies. Thus, variations in K,, may primarily reflect differences in the chemical nature of the organic matter. Using data from numerous research groups, Gerstl (1990) also examined the variability of log Kim values for 13 other nonionic compounds. He found the Kioc observations to be log normally distributed and to exhibit relative standard deviations for log Kioc values of about k lo- f 0.3 log units. An example is the herbicide atrazine, for which more than 200 observations were compiled (Fig. 9.9). DDT and lindane, two apolar compounds, exhibited similar variability in their log Kioc values as did atrazine. The variations can be attributed to the different methods applied by different groups and the variability in the


Figure 9.9 Frequency diagram illustrating the variability in the log Kioc values determined for atrazine for 217 different soil and sediment samples. The numbers on the X-axis indicate the center of a log K,,, range in which a certain number of experimental K,,, values fall. Data compiled by Gerstl (1990).

Atrazine

40 I

1.3 1.7 2.1 2.5 2.9 3.3 3.7

log K,,,/ (L - kg-loc)

qualitative nature of the organic matter in the wide range of soils and sediments used. In sum, careful determinations of nonionic organic compound absorption into natural organic matter appear to yield log K,,, values to about ? 0.3 log units (-+ l o )

1, AN I 'N I precision.

H H

atrazine Structural Characteristics of POM Relevant to Sorption

Let us now consider what the organic materials in soil and sediment sorbents are. As has become evident from numerous studies (see e.g., Thurman, 1985; Schulten and Schnitzer, 1997; Hayes, 1998), the natural organic matter present in soils, sediments, groundwaters, surface waters, atmospheric aerosols, and in wastewaters may include recognizable biochemicals like proteins, nucleic acids, lipids, cellulose, and lignin. But also, these environmental media contain a menagerie of macromolecular residues due to diagenesis (the reactions of partial degradation, rearrangement, and recombination of the original molecules formed in biogenesis). Naturally, the structure of such altered materials will depend on the ingredients supplied by the particular organisms living in or near the environment of interest. Moreover, the residues will tend to be structurally randomized. For example, soil scientists have deduced that the recalcitrant remains of woody terrestrial plants make up a major portion of the natural organic matter in soils (e.g., Schulten and Schnitzer, 1997). Such materials also make up an important fraction of organic matter suspended in freshwaters and deposited in associated sediments. Similarly, marine chemists have found that the natural organic matter, suspended in the oceans at sites far from land, consists of altered biomolecules such as polysaccharides and lipids that derived from the plankton and were subsequently modified in the environment (Aluwihare et al., 1997; Aluwihare and Repeta, 1999). At intermediate locales, such as large lakes and estuaries, the natural organic material in sediments and suspended in water appears to derive from a variable mixture of terrestrial organism and aquatic organism remains. An often-studied subset of these altered complex organic substances are commonly referred to as humic substances if they are soluble or


extractable in aqueous base (and insoluble in organic solvents), and humin or kerogen if they are not. The humic substances are further subdivided intofulvic acids if they are soluble in both acidic and basic solutions and humic acids if they are not soluble at pH 2. For a detailed overview of the present knowledge of humic materials, we refer to the literature (e.g., Hayes and Wilson, 1997; Davies and Gabbour, 1998; Huang et al., 1998; Piccolo and Conte, 2000). Here, we address only the most important structural features that are relevant to sorption of organic pollutants.

First, we note that natural organic matter that potentially acts as a sorbent occurs in a very broad spectrum of molecular sizes from the small proteins and fulvic acids of about 1 kDa to the huge complexes of solid wood and kerogen (>> 1000 kDa). Furthermore, natural organic matter is somewhat polar in that it contains numerous oxygen-containing functional groups including carboxy-, phenoxy-, hydroxy-, and carbonyl-substituents (Fig. 9.10). Depending on the type of organic material considered, the number of such polar groups may vary quite significantly. For example, highly polar fulvic acids may have oxygen-to-carbon mole ratios (OK ratios) of near 0.5 (Table 9.1). More mature organic matter (i.e., organic matter that has been exposed for longer time to higher pressures and temperatures in buried sediments) have O K ratios around 0.2 to 0.3, and these evolve toward coal values below 0.1 (Brownlow, 1979). These polar groups may become involved in H- bonding, which may significantly affect the three-dimensional arrangements and water content of these macromolecular media. Since many of the polar groups are acidic (e.g., carboxylic acid groups, phenolic groups) and because they undergo complexation with metal ions (e.g., Ca2+, Fe3+, A13+), pH and ionic strength have some impact on the tendency for the natural organic matter to be physically extended (when charged groups repulse one another) or coiled and forming domains that are not exposed to outside aqueous solutions. This may be particularly important in the case of “dissolved” organic matter (see Section 9.4).

In summary, we can visualize the natural organic matter as a complex mixture of macromolecules derived from the remains of organisms and modified after their release to the environment through the processes of diagenesis. This organic matter exhibits hydrophobic and hydrophilic domains. There is some evidence that the aggregate state of the organic matter may include portions with both fluid and rigid character. Borrowing terms commonly used in polymer chemistry, the inferred fluid domains have been referred to as “rubbery,” and the more rigid ones as “glassy” domains (Leboeuf and Weber, 1997; Xing and Pignatello, 1997). Other nomen- clature uses the terms soft and hard carbon, respectively (Weber et al., 1992; Luthy et al., 1997b). The glassy domain may contain nanopores (i.e., microvoids of a few nanometers size) that are accessible only by (slow) diffusion through the solid phase (Xing and Pignatello, 1997; Aochi and Farmer, 1997; Xia and Ball, 1999; Cornelissen et al., 2000). This would result in slow sorption kinetics (Pignatello and Xing, 1996). Thus, the natural organic matter may include a diverse array of compositions, resulting in both hydrophobic and hydrophilic domains, and formed into both flexible and rigid subvolumes. This picture suggests nonionic organic compounds may both absorb into flexible organic matter and any voids of rigid portions, as well as adsorb onto any rigid organic surfaces.


Figure 9.10 (a) Schematic soil humic acid structure proposed by Schulten and Schnitzer (1 997). Note that the "-" symbols stand for a linkages in the macromolecules to more of the same types of structure. (b) Schematic seawater humic substances structure proposed by Zafiriou et al. (1984). (c) Sche- matic black carbon structure proposed by Sergides et al. (1987).

0

0

0

Tab

le 9

.1 P

rope

rtie

s of

Org

anic

Com

pone

nts

that

May

Act

as

Sorb

ents

of

Org

anic

Com

poun

ds in

the

Env

iron

men

t

Com

pone

nt

Mol

e R

atio

M

olec

ular

Mas

s C

H

N

0

aver

age

(u) a

%

Aro

mat

icity

R

efer

ence

Bio

geni

c M

olec

ules

Prot

eins

Col

lage

n (p

rote

in)

Cel

lulo

se (

poly

sacc

hari

de)

Chi

tin (

poly

sacc

hari

de)

Lig

nin

(alk

alin

e ex

trac

t)

Lig

nin

(org

. sol

vent

ext

ract

)

Dia

gene

tic

Mat

eria

ls

Fulv

ic a

cids

so

il le

acha

te

brow

n la

ke w

ater

ri

ver

wat

er

grou

ndw

ater

Su

wan

nee

Riv

er fu

lvic

aci

d

brow

n la

ke w

ater

ri

ver

wat

er

“ave

rage

” so

il A

ldri

ch

Suw

anne

e R

iver

Hum

ic a

cids

Hum

in

Ker

ogen

Com

bust

ion-

Der

ived

Mat

eria

ls

NIS

T d

iese

l soo

t B

C f

rom

Bos

ton

Har

bor

sedi

men

t

1 .o

1 .o

1 .o

1 .o

1 .o 1 .o

1 .o

1 .o

1 .o

1 .o

1 .o 1 .o

1 .o

1 .o

1 .o

1 .o

1 .o

1 .o

1 .o

1 .o

1.6

1.7

1.7

1.8

1.1

0.98

1.04

0.

88

1.62

1.

04

0.87

0.80

1.

48

1.15

0.

78

0.94

1.

9 0.

4 to

1

0.1

1 .o

0.4

0.2

0.19

0.

31

< 0

.01

0.84

0.

13

0.64

<

0.01

0.

40

< 0.

01

0.33

< 0

.1

0.53

<

0.1

0.55

<

0.1

1.

09

< 0

.1

0.51

0.

1 0.

53

< 0

.1

0.54

<

0.1

0.

91

0.07

0

SO

0.

01

0.44

0.

02

0.61

0.

5 1.

1 0.

05 t

o 0.

3

0.01

6 0.

07

ca. l

o6 (c

otto

n)

2000

7800

90

00

3200

< 10

< 1

0 <

10

< 1

0 28

34

36

35

29

24

25

40

38

41

42

-100

1 2 2 2 2 2 3 3 4 3 7 3 4 5 6,7

67

7 8 9 10

Y

n

m

10

6 z

Mas

s av

erag

e.

Ref

eren

ces:

1. O

ser (

1965

); 2

. Xin

g et

al.

(199

4); 3

. Hai

tzer

et a

l. (1

999)

; 4. Z

hou

et a

l. (1

995)

; 5. S

chul

ten

and

Schn

itzer

(199

7);

6. A

rnol

d et

al.

(199

8); 7. C

hin

et a

l. (1

994)

; 8. G

arba

rini

and

Lio

n (1

986)

; 9. B

row

nlow

(197

9), 1

0. A

ccar

di-D

ey a

nd G

schw

end

(200

2).

BC

=bl

ack

carb

on.

W

h)

W 4


In addition to the natural organic matter present due to biogenesis and diagenesis, other identifiable organic sorbents, mostly derived from human activities, can be present (and would be included in an Ac measurement). Examples include combustion byproducts (soots and fly ash), plastics and rubbers, wood, and nonaqueous-phase liquids. The most potent among these other sorbents are various forms of black carbon (BC). Black carbon involves the residues from incomplete combustion processes (Goldberg, 1985). The myriad existing descriptors of these materials (soot, smoke, black carbon, carbon black, charcoal, spheroidal carbonaceous particles, elemental carbon, graphitic carbon, charred particles, high-surface- area carbonaceous material) reflect either the formation processes or the operational techniques employed for their characterization. BC particles are ubiquitous in sediments and soils, often contributing 1 to 10% of the&, (Gustafsson and Gschwend, 1998). Such particles can be quite porous and have a rather apolar and aromatic surface (Table 9.1). Consequently, they exhibit a high affinity for many organic pollutants, particularly for planar aromatic compounds. Therefore significantly higher apparent Ki,, values may be observed in the field as compared to values that would be predicted from simple partitioning models (Gustafsson et al., 1997; Naes et al., 1998; Kleineidam et al. 1999; Karapanagioti et al., 2000).

Another example involves wood chips or sawdust used as fills at industrial sites. Wood is also a significant component of solid waste, accounting for up to 25 wt% of materials at landfills that accept demolition wastes (Niessen, 1977). Wood is composed primarily of three polymeric components: lignin (25-30% of softwood mass), cellulose (40-45% of softwood mass), and hemicellulose (remaining mass) (Thomson, 1996). As has been shown by Severton and Banerjee (1996) and Mackay and Gschwend (2000), sorption of hydrophobic organic compounds by wood is primarily controlled by sorption to the lignin. This is not too surprising when considering the rather polar character of cellulose and hemicellulose as compared to lignin (compare Q/C and H/C ratios in Table 9.1). Also synthetic polymers such as polyethylene (Barrer and Fergusson, 1958; Rogers et al., 1960; Flynn, 1982; Doong and Ho, 1992; Aminabhavi and Naik, 1998), PVC (Xiao et al., 1997), and rubber (Barrer and Fergusson, 1958, Kim et al., 1997) and many others are well known to absorb nonionic organic compounds. If such materials are present in a soil, sediment, or waste of interest, then they will serve as part of the organic sorbent mix. Finally, a special organic sorbent that may be of importance, particularly, when dealing with contamination in the subsurface, is nonaqueous phase liquids (NAPLs, Hunt et al., 1988; Mackay and Cherry, 1989). These liquids may be immobilized in porous media and serve as absorbents for passing nonionic organic compounds (Mackay et al., 1996). In such cases we may apply partition coefficients as discussed in Section 7.5 (Eq. 7-22) to describe sorption equilibrium, but we have to keep in mind that the chemical composition of the absorbing NAPL will evolve with time.

In conclusion, sorption of neutral organic chemicals to the organic matter present in a given environmental system may involve partitioning into, as well as adsorption onto, a variety of different organic phases. Thus, in general, we cannot expect linear isotherms over the whole concentration range, and we should be aware that predictions of overall K,,, values may have rather large errors if some of the important organic materials present are not recognized (Kleineidam et al., 1999).


Conversely, with appropriate site-specific information, reasonable estimates of the magnitude of sorption coefficients can be made (see below).

Determination of Kioc Values and Availability of Experimental Data

K,, values are available for a large number of chemicals in the literature. The vast majority of these Kioc's have been determined in batch experiments in which a defined volume of water is mixed with a given amount of sorbent, the resultant slurry is spiked with a given amount of sorbing compound(s), and then the system is equilibrated with shaking or stirring. After equilibrium is established, the solid and aqueous phases are mostly separated by centrifugation or filtration. In most studies, only the aqueous phase is then analyzed for the partitioning substance, and its concentration in the solid phase is calculated by the difference between the total mass added and the measured mass in the water. Direct determinations of solid phase concentrations are usually only performed to verify that other loss mechanisms did not remove the compound fi-om the aqueous phase (e.g., due to volatilization, adsorption to the vessel, and/or degradation). Kioc is then calculated by dividing the experimentally determined Kjd (= Cis / C,) value by the fraction of organic carbon, A,, of the sorbent investigated (Eq. 9-22).

Of course, a meaningful Kioc value is obtained only if sorption to the natural organic material is the dominant process. This may be particularly problematic for sorbents exhibiting very low organic carbon contents. Also, solid-water contact times are sometimes too short to allow sorbates to reach all the sorption sites that are accessible only by slow diffusion (Xing and Pignatello, 1997); thus, assuming sorptive equilibrium may not be appropriate. This kinetic problem can be especially problematic for equilibrations that employ sorbate solutions flowing through columns containing the solids. Finally, errors may be introduced due to incomplete phase separations causing the presence of water (containing dissolved compound) with the solid phase, as well as colloids (containing sorbed compound) in the aqueous phase. Hence, the experimentally determined apparent solid-water distribution coefficient, Kt:", is not equal to the "true" K;d but is given by:

(9-23)

where Cis is the compound concentration on the separated particles (mol . kg-' solid)

C;, is the compound concentration in the water (mol . L-')

V,, is the volume of water left with the separated particles (L . kg-' solid)

CiDoc is the compound concentration associated with colloids (mol . kg-' oc)

[DOC] is the concentration of organic matter in the colloids (expressed as C) remaining with the bulk water (kg oc . L-')

By dividing the numerator and denominator of Eq. 9-23 by Ciw, and then substituting C,/Ciw by Kid and CiDoc/Ciw by KiDOC, we may rewrite Eq. 9-23 as:


(9-24)

The expression indicates that the apparent solid-water distribution coefficient will equal the “true” one only if V,, (( K;d and if KiDoc . [DOC] (( 1. For weakly sorbing compounds (low Kid), this equation suggests that the experimental Kl~pp, and thus Kzp , may be erroneously high. For compounds that do tend to sorb (high K,,,) and in situations where organic colloids are substantial (high [DOC]), batch observations of solid-water partitioning produce lower distribution coefficients than Kid. Note that these phase separation difficulties are probably one of the major explanations for the so-called “solids concentration effect” in which Kid appears to decrease with greater and greater loads of total solids and thus DOM colloids in batch sorption systems (Gschwend and Wu, 1985). Note also that these problems may also be important for other colloid-containing systems such as where sorption to clay minerals plays a major role (see Chapter 11). Finally, particularly in older studies, radiolabeled chemicals of poor purity were used, and this can also have an influence on the result (Gu et al., 1995).

Considering all these experimental problems, as well as the natural variability of natural organic sorbents, it should not be surprising that Kjo, values reported in the literature for a given chemical may vary by up to an order of magnitude or even more. This is particularly true for polar compounds for which uncontrolled solution conditions like pH and ionic strength may also play an important role. Thus, when selecting a K;,, value from the literature, one should be cautious. In this context, it should be noted that Ki,, values are log-normally distributed (normal distribution of the corresponding free energy values), and therefore log Ki,, values, not Kj,, values, should be averaged when several different KjOc7s have been determined (Gerstl, 1990).

For the following discussions, we will primarily use Kio, values from compilations published by Sabljic et al. (1995) and Poole and Poole (1999). According to these authors, the values should be representative for POM-water absorption (i.e., they have been derived from the linear part of the isotherms). Furthermore, many of the reported Kim’s are average values derived from data reported by different authors. Distinction between different sources of sorbents (e.g., soils, aquifer materials, freshwater, or marine sediments) has not been made. Nevertheless, at least for the apolar and weakly monopolar compounds, these values should be reasonably representative for partitioning to soil and sediment organic matter.

Estimation of Kjoc values

Any attempt to estimate a K,, value for a compound of interest (with its particular abilities to participate in different intermolecular interactions) should take into account the structural properties of the POM present in the system considered. To this end, the use of multiparameter LFERs such as the one that we have applied for description of organic solvent-water partitioning (Eq. 7-9) would be highly desirable (Poole and Poole, 1999). Unfortunately, the available data do not allow such analyses, largely due to the very diverse solid phase sources from which reported Kioc values have been derived.


Figure 9.11 Plot of log K,,, versus log KiOw for PAHs (+) and for a series of alkylated and chlorinated benzenes and biphenyls (PCBs) (A). The slopes and intercepts of the linear regression lines are given in Table 9.2.

I

h

0 6 - 0 7

b . 5 - d.

8 4 - k-

Y

--.

rn 0

3 - -

2 -

chlorinated benzenes, PCBs

1 I I

Therefore, for estimates of K,,,'s it is more feasible to use compound class-specific LFERs. These include correlations of log K,,, with molecular connectivity indices (or topological indices; for an overview see Gawlik et al., 1997), with log C;G'(L) (analogous to Eq. 7-1 l), and with log K,,,. Although molecular connectivity indices or topological indices have the advantage that they can be derived directly from the structure of a chemical, they are more complicated to use and do not really yield much better results than simpler one-parameter LFERs using C;$t(L) or K,,, as compound descriptors.

Since C:$t (L) and Ki,, can be related to each other (Eq. 7-1 1, Table 7.3), here we will confine ourselves to log Ki,, - log Ki,, relationships. Table 9.2 summarizes the slopes a and intercepts b derived for some sets of organic compounds by fitting:

log Ki,, = a. log Ki,, + b (9-25)

Note that the Ki,, values used tend to represent mostly absorption into soil and sediment POM. Therefore, any estimates using equations such as the ones given in Table 9.2 should be considered to be within a factor of 2 to 3. Furthermore, such LFERs should be applied very cautiously outside the log K,, - log Ki,, range for which they have been established. This is particularly critical for LFERs that have been derived only for a relatively narrow log Ki,, range (e.g., the phenyl ureas).

Let us make some general comments on this type of LFER. First, reasonable correlations are found for sets of compounds that undergo primarily London dispersive interactions (Fig. 9.11 ; alkylated and chlorinated benzenes, chlorinated biphenyls). Good correlations are also found for sets of compounds in which polar interactions change proportionally with size (PAHs) or remain approximately


Figure 9.12 Plot of log K,,,, versus log K,,, for a alkylated and halogenated (R, = alkyl, halogen) phenylureas (R2 = R, = H; A, halogen, see margin below), phenyl-methylureas (R2 = CH3, R3 =

H, n), and phenyl-dimethylureas (R2 = R3 = CH3, 0). The slope and intercept of the linear regression using all the data is given in Table 9.2 (Eq. 9-261); each subset of ureas would yield a tighter correlation if considered alone (e.g., Eq. 9-26j).

h

8 2.5 b

d . 2

7

Y

\ 0

kQ 1.5 -

1

0.5 0.5 1 1.5 2 2.5 3 3.5

log Kj ow

constant (chlorinated phenols). These results are reasonable based on our previous discussions of organic solvent-water partitioning (Chapter 7).

It should also be not too surprising that poorer results are obtained when trying to correlate sets of compounds with members exhibiting significantly different H-acceptor and/or H-donor properties. This is the case for the halogenated C,-, C2-, and C,-compounds. Combining the entire set leads to an R2 of only 0.68 (Eq. 9-26d). Focusing on the chloroalkenes, the correlation is much stronger (R2 of 0.97 although N is only 4); while for the polyhalogenated alkanes with and without bromine correlations are much more variable. This can be understood if we recall that compounds like CH,C12 exhibit H-donor and H-acceptor capabilities (e.g., for CHC1, a, = 0.10 and pi = 0.05) while C1,CCH3 has only H-acceptor ability (ai = 0, p, = 0.09) and CC1, has neither (Table 4.3). Hence, lumping such sets of compounds in a single-parameter LFER should yield variability as is seen. Such H-bonding variability also occurs within the large set of phenyl ureas that are used primarily as herbicides (some with -NH2, others with -NH-CH,, and finally some with -N(CH,),). In the case of the phenyl ureas, a significantly better correlation can be obtained for any subset of these compounds exhibiting consistent H-bonding on the terminal amino group (Fig. 9.12). Consequently, more highly correlated relationships with a single parameter like log Kio, are also found for these subsets (Table 9.2). These examples demonstrate that care has to be taken when selecting a set of compounds for the establishment of one-parameter LFERs. Hence, published LFERs relating log Kio, values to log K,, or related parameters (liquid aqueous solubilities or chromatographic retention times; see Gawlik et al., 1997, for review) should be checked to see that the “training set” of sorbing compounds have chemical structures that ensure that they participate in the same intermolecular interactions into the two partitioning media.

r ? 0

Tab

le 9

.2 L

FER

s R

elat

ing

Part

icul

ate

Org

anic

Mat

ter-

Wat

er P

artit

ion

Coe

ffic

ient

s an

d O

ctan

ol-W

ater

Par

titio

n C

onst

ants

at 2

0 to

25

°C f

or S

ome

Set

s of

Neu

tral

Org

anic

Com

poun

ds: S

lope

s an

d In

terc

epts

of

Eq.

9-2

5 a

Equ

atio

n Se

t of

Com

poun

ds

~~

9-26

a A

lkyl

ated

and

chl

orin

ated

ben

zene

s, P

CB

s (k

apo

lar)

0.

74

0.15

2.

2 to

7.3

0.

96

32

9-26

1,

PAH

s (m

onop

olar

) 0.

98

-0.3

2 2.

2 to

6.4

0.

98

14

9-26

c C

hlor

inat

ed p

heno

ls (n

eutr

al sp

ecie

s; b

ipol

ar)

0.89

-0

.15

2.2

to 5

.3

0.97

10

9-26

d C

1 - an

d C

2-ha

loca

rbon

s (ap

olar

, mon

opol

ar, a

nd b

ipol

ar)

0.57

0.

66

1.4

to 2

.9

0.68

! 19

9-26

e O

nly

chlo

roal

kane

s (m

ix)

0.42

0.

93

0.59

! 9

9-26

f O

nly

chlo

roal

kene

s (k

apo

lar)

0.

96

-0.2

3 0.

97

4

9-26

g O

nly

com

poun

ds in

clud

ing

brom

ine

(mix

) 0.

50

0.8

1 0.

49!

6

9-26

h A

ll ph

enyl

urea

s (b

ipol

ar)

0.49

1.

05

0.5

to 4

.2

0.62

! 52

9-26

i O

nly

alky

late

d an

d ha

loge

nate

d ph

enyl

urea

s, p

heny

l-

0.59

0.

78

0.8

to 2

.9

0.87

27

9-26

j O

nly

alky

late

d an

d ha

loge

nate

d ph

enyl

urea

s (b

ipol

ar)

0.62

0.

84

0.8

to 2

.8

0.98

13

a D

ata

from

Sab

ljic

et a

l. (1

995)

, Chi

ou e

t al

. (1

998)

, and

Poo

le a

nd P

oole

(19

99).

The

dat

a fo

r ch

lori

nate

d ph

enol

s ha

ve b

een

take

n in

par

t fr

om S

chel

lenb

erg

et a

l. (1

984)

. ' K

, in

L.k

g-'o

c.

See

Fig

. 9.

12,

all

met

hylu

reas

, and

phe

nyl-

dim

ethy

lure

a (bi

pola

r)

Ran

ge o

f ex

peri

men

tal

valu

es f

or w

hich

LFE

R h

as b

een

esta

blis

hed.

N

umbe

r of

com

poun

ds u

sed

for

LFE

R.

com

poun

ds. Y Se

e Fi

g. 9

.12,

onl

y A

.

v, 3 g 5

s h

2

U z W w

0

w


Kioe as a Function of Sorbate Concentration

Let us now come back to the issue of linearity of the isotherm and dependency of Kid on the sorbate concentration. In numerous field studies in which both particle- associated and dissolved concentrations of PAHs are measured, apparent Ki,, values are up to two orders of magnitude higher than one would have predicted from a simple absorption model (Gustafsson and Gschwend, 1999). If a natural soil or sediment matrix includes impenetrable hydrophobic solids on which the chemical of interest may adsorb, the overall Kioc value must reflect both absorption into recent natural organic matter and adsorption onto these surfaces.

We start out by considering the effect of such adsorption sites on the isotherms of apolar and weakly monopolar compounds. For these types of sorbates, hydrophobic organic surfaces and/or nanopores of carbonaceous materials are the most likely sites of adsorption. Such hydrophobic surfaces may be present due to the inclusion of particles like coal dust, soots, or highly metamorphosed organic matter (e.g., kerogen). Because of the highly planar aromatic surfaces of these particular materials, it is reasonable to assume that planar hydrophobic sorbates that can maximize the molecular contact with these surfaces should exhibit higher affinities, as compared to other nonplanar compounds of similar hydrophobicity.

Let us evaluate some experimental data. To this end, we use a dual-mode model (Eq. 9-6). This model is a combination of a linear absorption (to represent the sorbate’s mixing into natural organic matter) and a Freundlich equation (as seen €or adsorption to hydrophobic surfaces or pores of solids like activated carbons):

The value of the partition coefficient, Kip, is given by the product,foci(ioc, wheref,, and Ki,, apply only to the natural organic matter into which the sorbate can penetrate. The value of KiF is less well understood, but recent observations suggest it should be related to the quantity of adsorbent present (e.g., the fraction of “black carbon” in a solid matrix, fb,) and the particular compound’s black-carbon- normalized adsorption coefficient (e.g., KibC). Typical values of the Freundlich exponent are near 0.7. Hence, in a first approximation the data should fit:

(9-27)

Observations certainly fit this type of dual-mechanism model. For example, Xia and Ball (1999) recently examined the sorption of several organic compounds to an aquifer sediment. They measured that sediment’s&, to be 0.015. Using a literature value of the Kpyreneoc of 104.7 (Gawlik et al., 1997), it is clear that the pyrene sorption they observed greatly exceeded expectations based on only&, times Kpyrene oc (Figure 9-13a). Subtracting this absorption contribution to the total Kpyrene ,, and using a recently reported value for Kpyrenebc of 106.5 (Bucheli and Gustafsson, 2000), the data indicate&, in this aquifer sediment contributed about 0.6% of the solid mass (a large fraction of that Miocene sediment’s remaining reduced carbon content). Using this

f b c , the entire pyrene sorption isotherm was well fit using Eq. 9-27 (Figure 9- 13a). Moreover, fixing& at 0.006, the isotherms for the other sorbates tested by Xia and


and Accardi-Dey and Gschwend 9 (2001). 2 -

(b) Holding&, at 0.006, values of KibC can be estimated for all the 1

( 4

1 2 3 4 5 6 7 other sorbates (all planar) tested on -1 1 10 100

this aquifer sediment: benzene, log ciw / (yg.L-’) log Kiow chlorobenzene, 1,2-dichlorobenzene, naphthalene, 1,2,4-trichloro-

chlorobenzene, phenanthrene, and pyrene.

fluorene* ,2,475-tetra- Ball can be used to extract those compounds’ Kibc values in [(pg . kg-’ bc)(pg . L-’)-“i]. These data suggest that for planar sorbates a value of Kibc can be estimated via:

log Kibc G 1.6 log Ki,, - 1.4 ( N = 9, R2 = 0.98) (9-28)

Consistent with experience with adsorbents like activated carbons, the fitted KIbc values are greater for sorbates with larger hydrophobicities (Fig. 9.13b). Note that when using Freundlich isotherms, the KiF value depends nonlinearly on the units in which the concentration in the aqueous phase is expressed (see Problem 9.5).

Neglecting the contribution of adsorption, especially for planar compounds and at low concentrations, may cause substantial underestimation of Kid. This is shown in Illustrative Example 9.2 for phenanthrene sorption to various soils and sediments (Huang et al., 1997).

Illustrative Example 9.2 Evaluating the Concentration Dependence of Sorption of Phenanthrene to Soil and Sediment POM

Huang et al. (1997) measured sorption isotherms for phenanthrene on 21 soils and sediments. All isotherms were nonlinear with Freundlich exponents ni (Eq. 9-1) between 0.65 and 0.9. For example, for a topsoil (Chelsea I) and for a lake sediment (EPA-23), interpolating the isotherm data yields the following “observed” sorbed concentrations, Cis, in equilibrium with dissolved concentrations, C,, of 1 pg . L-’ and 100 pg . L-I, respectively:

Ciw c i s

(Pug * L-I) (pg . kg-’ solid) Chelsea-I EPA-23

1 3,200 1,700 I00 9 1,000 5 1,000


i - phenanthrene

log K,,, = 4.57 log K,, = 4.3

Problem

Using Eq. 9-27, estimate the equilibrium solid phase concentrations, CIS, of phenanthrene for this topsoil and this sediment for aqueous concentrations, C,,, of 1 and 100 yg.L-'. Compare these values with the concentrations obtained from interpolation of the sorption isotherms (see above):

Answer

For this nonionic, planar compound, you have to take into account both absorption to POM and adsorption to a high-affinity sorbent (e.g., black carbon). For Chelsea I soil, fOc was measured as 0.056 kg oc . kg-' solid. The fbc was not measured, but in sediment samples it is typically between I and 10% of thef,, (Gustafsson and Gsch- wend, 1998). Use the full range of 1 to 10% to see the possible impact of adsorption to black carbon (ie., fbc = 0.00056 to 0.0056 kg bc.kg-' solid). Assume n, = 0.7 and use Eq. 9-28 to estimate Klbc:

log&,,= 1.6 10gKjo,- 1.4=(1.6)(4.57))- 1.4=5.9

Insertion of&,, Ki,,,fbc, and KibC into Eq. 9-27 yields Cis values for Ciw=l and 100 yg . L-', respectively:

for C,, = 1 pg.L-':

c, = (0.056)( 104.3)( 1) + (0.00056 to 0.0056)( 1)0.7 = 1100 + (440 to 4400) = 1540 to 5500 y g . kg-' solid

(observed 3200 pg . kg-' solid)

Note that a calculation based only on the product, fO&,, would underestimate the oberved values by about a factor of three.

for C,, = 100 yg.L-':

Cis = (0.056)( 104.3)( 100) + (0.00056 to 0.0056)( 105.9)( = 110,000 + (1 1,000 to 1 10,000) = 121,000 to 220,000 yg . kg-' solid

(observed 9 1,000 pg . kg-' solid)

In this case, the estimate based only onfo&ioc is very close to the experimental result, indicating that for high substrate concentrations partitioning into POM is the dominant sorption mechanism.

For EPA-23 lake sediment, f,, was measured as 0.026 kg oc.kg-' solid. Again assuming the same K,,, and K,, values for phenanthrene and the same&, range, one obtains:

for Ci, = 1 pg . L-' :

Cis = 520 + (210 to 2100) = 730 to 2600 pg. kg-' solid (observed 1700 pg . kg-' solid)

Figure 1 Predicted versus experimental sorbed phenanthrene concentrations for a series of soils and sediments. The diamonds indicate solids equilibrated with 1 pg . L-'; squares are for solids equilibrated with 100 pg . L-I.


7 ,

+ 1 pg-L-' T

b I 0 100 pg.L-1 Y

+

n I I I I I I " 0 1 2 3 4 5 6 7

predicted log Cphenanthrene / @g.kg-' solid)

and:

for Ci, = 1 OO pg . L-' :

Cis = 52,000 + (5200 + 52,000) = 57,000 to 104,000 pg . kg-' solid (observed 5 1,000 pg . kg-' solid)

Hence, as for the Chelsea I topsoil,f,, .Kioc underestimates the observed sorption by about a factor of 3 at Ci, = 1 pg . L-', whereas at 100 pg . L-', sorption is dominated by POM-water partitioning.

Predictions of sorbed phenanthrene concentrations equilibrated with 1 or 100 pg . L-' dissolved concentrations for all the soils and sediments (assuming f , = 0.05 f,,) examined by Huang et al. (1997) are shown in Fig. 1. Values for 100 pg.L-' fall somewhat below the extrapolation from 1 yg . L-' observations, indicating the shift from adsorption to absorption as the dominating mechanism.

For compounds other than PAHs, unfortunately there are not enough data available that would allow a more general analysis of the concentration dependence of K,, values. Nevertheless, a few additional observations may give us some better feeling of the magnitude of this dependence. For example, for sorption of smaller apolar and weakly monopolar compounds (e.g., benzene, chlorobenzene, 1,2-dichIorobenzene, tetrachloroethene, dibromoethane) to soil (Chiou and Kile, 1998) or aquifer materials (Xia and Ball, 1999 and 2000), not more than a factor of 2 difference in K,, was found between low and high sorbate concentrations. A somewhat more pronounced effect (i.e., factor 2 to 3) was observed for sorption of the more polar


7 y 3 herbicides atrazine (Sposito et al., 1996; for structure see Fig. 9.9) and diuron

q N \ < CH3 (Spurlock and Biggar, 1 9 9 4 ~ ; for structure, see margin) as well as for 3,5-dichlorophenol to soil organic carbon (Chiou and Kile, 1998). Note again that in the case of polar compounds, other high-affinity sites present in POM may cause the CI

CI diuron nonlinear sorption behavior.

In summary, we should realize that, when applying Kioc values determined at high concentrations or derived from LFERs such as Eq. 9-26 (Table 9.2), we may underestimate equilibrium sorption at low concentrations (i.e., below 1% of the compound’s solubility) by a factor of 2 or more. Due to competition with other sorbates present in a natural system, the effect of specific adsorption could, however, be significantly attenuated (Chiou et al., 2000). Furthermore, the abundance of specific adsorption sites may be rather low in certain environments. Hence, in cases in which the effect can be expected to be moderate and in which we need only to get an order of magnitude estimate of K,,, we may decide to neglect the nonlinearity of the isotherm. In other situations, however, such as the PAH case discussed in Illustrative Example 9.3, adsorption to carbonaceous materials or other high-affinity sorbents present in significant abundances has to be taken into account.

Illustrative Example 9.3 Estimating Pore Water Concentrations in a Polluted Sediment

Commonly, we are interested in estimating the concentration of a chemical in the pore water of a sediment bed. For example, we may be concerned that a pollutant is diffusing out of a contaminated sediment bed into the overlying waters; the rate of this exchange will be proportional to the difference in dissolved concentrations in the pore water and the overlying water (see Chapter 19). Or we may be interested in the biouptake of the pollutants by invertebrates living in the bed (see Chapter lo). Such uptake may be proportional to the pore water concentrations, which themselves are proportional to the concentrations in the solids.

Problem

In a sediment core from Boston Harbor (off Spectacle Island, Massachussetts, USA), McGroddy et al. (1 996) measured sediment and pore water concentrations of pyrene and 2,2’,4,5,5’-pentachlorobiphenyl (PCB101). For a set of depths, she obtained the following values given below. Estimate the “mean” pore water concentrations of these two chemicals. Assume that& is 10% off,, in this area (Gustafsson and Gschwend, 1998). Neglect sorption of PCBl 01 to black carbon since this is not a planar molecule.

Answer

Rearranging Eq. 9-27 and solving for Cipore water (= Ciw) yields:

Note that because of the nonlinearity of the Freundlich term, you may not isolate


Depth interval "6, Concentration in sediment

surface) Pyrene PCBlOl (cm below (kg oc . kg-' solid) C,,/(pg. kg-' solid) @ / /

pyrene

lOgK,,,=513 6-8 0.035 1900 15 log K,bc = 6.50

K,bc in fJJg.kg-ibc)fJJg.Ll)-0.7 (Bucheli and Gustafsson, 2000)

8-10

10-12

0.034

0.030

1700 15

22,30 2600,1600

means 0.032 2000 20 CI, C',

Cip,, water. Therefore, to solve Eq. 1, you will have to guess the solution and revise your guess until you satisfy the equation.

Estimate the K,,, values of the compounds from Ki,, (Eqs. 9-26b and 9-26a, respectively):

CI CI

2,2'.4.5,5-pentachlorobiphenyl (PCB101)

log K,,, = 6 36

pyrene: log Kioc= (0.98)(5.13)-0.32 = 4.7

PCB101: log K,, = (0.74)(6.36)+0.15 = 4.9

Insertion of Cis = 2000 pg . kg-' solid together withf,, = 0.032 and f b c = 0.1 f , , =

0.0032 and the appropriate Ki,, and Kibc values into Eq. 1 allows an estimate for pyrene via:

C,,, (pg . L-') = 2000/[(0.032)( 104.7) + 0.003)( 106.5)(Cipore ,,,J 0.3 ]

By trial and error, one may find Cjporewater G U.Zpg-L-', not too far from the measured value of 0.04 pg.L-' (McGroddy et al., 1996). Neglecting adsorption to black carbon would have yielded a much higher estimate:

Ciporewater= 2000/[(0.032)(104.7)] = 1.2 pg.L-'

For PCB 10 1, the pore water concentration at equilibrium is estimated to be:

Ciporewater = 20/[0.032)(104.9)] = 0.008pg.L-'

which is, again, quite similar to the measured value of 0.004 pg .L-' (McGroddy et al., 1996).


Effect of Temperature and Solution Composition on Kio,

Temperature. Like organic solvent-water systems (Eq. 7-4), over a narrow temperature range, we may express the temperature dependence of K,,, by:

A P O M ~ H ~ 1 R T

In Ki,, = - .- + const. (9-29)

with:

APOMWH~ = HgOM - H E (9-30)

where HIEoM and H; are the excess enthalpies of i in the sorbed phase (an average value) and in the aqueous phase, respectively. Previously (Table 5.3) we saw that for many small organic compounds, H: is relatively small. Significant negative deviations from zero can be expected primarily for small monopolar compounds (e.g., -20 to -30 kJ.mol-I) and large apolar compounds (e.g., +20 to 30 kJ.mo1-'). Note that HgoM represents the average excess enthalpy of the compound for the various sorption sites and may, therefore, also depend on the concentration range considered. Hence, at low concentrations where sorption to specific sites may dominate the overall K,,, (see above), HzOM may be different as compared to higher concentrations where partitioning is the major process. For sorption of apolar and weakly polar compounds to hydrophobic sites, including surfaces of carbonaceous materials, we may still assume that H,", is primarily determining A P O M ~ HI (Bucheli and Gustafsson, 2000). For polar compounds that may undergo H-bond interactions with specific sites present in POM, however, may change significantly with concentration. For example, for sorption of diuron and other phenylurea herbicides to soils, HgOM values were found to vary between about -40 to -50 kJ . mol-' at very low concentrations to f 10 kJ . mol-' at concentrations where nonspecific partitioning can be assumed to dominate the overall sorption (Spurlock, 1995). Since for these compounds, HE is about +10 kJ.mol-' (Spurlock, 1995), one obtains ALPOM~H, values of between -50 and -60 kJ. mol-' at very low concentrations, and between -20 and 0 kJ. mol-' for the high concentration range. Hence, as in many other cases, partition coefficients are rather insensitive to temperature, whereas adsorption coefficients increase by about a factor of 2 with a 10 degree decrease in temperature. Consequently, in such situations, one can expect that the nonlinearity of the isotherm increases with decreasing temperature.

p H and Ionic Strength. For sorption of apolar and weakly polar compounds to POM we may neglect the effect of pH and ionic strength on the sorption properties of the organic phase. However, particularly for higher salt concentrations, we may have to take into account the effect of salt on the activity coefficient of the compound in the aqueous phase. As discussed in Section 5.4 and applied to air-water partitioning in Chapter 6 (Eq. 6- 19), we may quantify this effect simply by:

(9-3 1)

K: values for seawater are given in Table 5.7 for a series of compounds. Using a typical value of 0.3 to 0.35 for PAHs and [salt],,, = 0.5 M for seawater, we can see that the presence of the salt will increase K,,, by a factor of about 1.5.


Organic Cosolvents. Sometimes we have to deal with the behavior of organic chemicals present in aqueous solutions containing an appreciable amount of an organic cosolvent (see Section 5.4). Let us examine the effect of such organic cosolvents on the Ki,, value of a given compound by looking at sorption to POM simply as a partitioning process between water and a complex organic solvent exhibiting a molar volume vK and a density p,. Hence, by analogy to Eq. 7-2 we express KiOc as:

(9-32) y i w . v w

yioc . v, . PIX Kim = -

We can now evaluate how a given organic cosolvent will affect the various parameters in Eq. 9-32. In Section 5.4 we discussed the dependence of the activity coefficient of a compound in a solvent-water mixture on the fraction of the cosolvent. We have seen that, depending on solute and cosolvent considered, this dependence may be quite complex (Figs. 5.6 and 5.7; Table 5.8). In the following discussion, we confine ourselves to rather small cosolvent concentrations (i.e.,f, < 0.2 to 0.3) for which we may assume a log-linear relationship (Eq. 5-32). We may then express the activity coefficient, yit, of compound i in the solvent-water mixture as:

Yi[(f") = yjw -lO-":.f' (5-33)

Inspection of Table 5.8 shows that for naphthalene, for a variety of organic cosolvents, 0: varies between 2.0 (for glycerol) and 6.5 (for acetone). Recall that 0: is referred to as cosolvency power. Hence, in a 20% cosolvent-80% water mixture (i.e., f , = 0.2), Yie will be lower by a factor of between 2.5 (glycerol) and 20 (acetone) as compared to "/iw. Considering this rather large effect, as a first approximation, we may neglect the change in the molar volume of the solvent-water mixture, particularly when dealing with cosolvents with relatively small v, values. For example, the molar volume of a 20% methanol ( v, = 0.0406 L .mol-') - 80% water mixture is 0.02 L.mo1-' as compared to 0.018 L.mol-' for pure water (see Illustra- tive Example 5.5). Thus, if the cosolvent has no effect on the properties of the POM sorbent, the effect of the cosolvent on Ki,, (or Kid), could be described simply by:

~ i , , s o l v / w = Ki, . IO-"'~' (9-33)

Note again that 0: depends both on the solute i and on the type of the cosolvent. Furthermore, o,C is not a constant and may vary depending on the&-range for which it has been determinated. In general, 0," will increase with increasingf, (e.g., Figs. 5.6 and 5.7).

Depending on the type of sorbate, sorbent, and cosolvent considered, we should not necessarily assume that the effect of the cosolvent on the sorption properties of the sorbent may be completely neglected. In order to account for this effect, the empirical model (Eq. 9-33) can be extended by introducing a second parameter a (Rao et al., 1985):

K*oc,so'"/w = Kj, . lo-".":.fv (9-34)

Note that a quantifies how the organic cosolvent changes the nature of the sorbent with respect to its quantity (e.g., by swelling and including cosolvent in the sorbent!)


and quality. Thus a will depend not only on the type of sorbent and cosolvent, but also on the type of compound (hence, we could also use a subscript i). Note that in the literature a is often assumed to be independent of the solute considered, which is reasonable when dealing with apolar compounds or structurally related compounds. Furthermore, because it is likely that the cosolvent has a different impact on the various different sorbents that may be present in a given environment (e.g., carbonaceous material versus humic components), a may vary withf, and even with solute concentrations. Nevertheless, at least for a “water-like” cosolvent such as methanol, the simple model Eq. 9-34 has been applied with reasonable success to fit experimental data, particularly for POM-water partitioning of nonpolar or weakly polar compounds. In most of the studies involving methanol as cosolvent, a has been found to be close to 1. That is, this cosolvent did not have a significant impact on the organic sorbent (Rao et al., 1990; Spurlock and Biggar, 1994b; Bouchard, 1998). An example calculation is given in Illustrative Example 9.4. For more details we refer to the literature (e.g., Nkedi-Kizza et al., 1985; Fu and Luthy, 1986; Rao et al., 1990; Wood et al., 1990; Lee et al., 1993; Kimble and Chin, 1994; Nzengung et al., 1996; Lee and Rao, 1996; Bouchard, 1998).

Illustrative Example 9.4 How Much Does the Presence of 20% Methanol in the “Aqueous” Phase Affect the Retardation of Phenanthrene in an Aquifer?

Problem

Consider the transport of phenanthrene in an aquifer exhibiting a porosity, 4 , of 0.2, and an average density, ps, of the aquifer material of 2.5 kg so1id.L-I. Furthermore, assume that the average organic carbon content of the aquifer material is 0.5% (i.e.,foc = 0.005 kg oc. kg-’ solid). Calculate the retardation factor Rf, (i.e., A;’. see Section 9.2) of phenanthrene in this aquifer if the groundwater consists (a) of pure water, (b) of a 20% methanol/80% water volume mixture, and (c) of a 30% methanol/70% water volume mixture.

Answer

The retardation factor R, is given by (Eq. 9-12):

i - phenanthrene

Inserting the values for p, and 4 yields:

R, = 1 + 10 (kg solid. L-I) f& (L . kg-l solid) (1)

Case (a) Assume that partitioning to the POM is the major sorption mechanism. Use an average Ki,, value of 2 x lo4 L . kg-’ oc to estimate Kid (Eq. 9-22):


Kid =f,. K,, = (0.005) (2 x lo4) = 100 L. kg-' solid

Insertion of this value into Eq. 1 above yields an R, value of about I000 if the groundwater consists of pure water.

Case (b) To estimate the effect of the methanol on Kid , assume that a is about 1 in Eq. 9-34. Get the olC value of phenanthrene from the PAH data summarized in Fig. 5.8 using (see figure caption):

log(^::^ Irs,"') = (0.0104) Kx -0.668 forfv,Me,, = 0.2

(see also Illustrative Example 5.5, Answer (b)).

In this case, 0' is given by (see also Section 5.4; Eqs. 5-30 and 5-31):

The Vix of phenanthrene calculated by the method described in Box 5.1 is 145.4 cm3 mol-', and, therefore:

o,C = 0.84 / 0.2 = 4.2

Insertion of this value into Eq. 9-35 with a = 1 yields:

Hence, the 20% methanol lowers the apparent Kioc by a factor of about 7 ; that is, R, z I40 instead of 1000 in the case of pure groundwater.

Case (c) Use theo,"value derived above for the 0 to 20% methanol range. Note that in reality,o; may be somewhat larger at this higher methanol fraction. Insertion of 0," = 4.2 into Eq. 9-35 with a = 1 andfY = 0.3 yields:

Ki0c,30% MeOH = 0.055 Kioc

which yields an R, value of 55.

3 14 Sorption I: Sorption Processes Involving Organic Matter

Sorption of Neutral Compounds to “Dissolved” Organic Matter (DOM)

“Dissolved” organic matter (DOM) is operationally defined as the organic material that passes a filter, commonly having pore size near 1 pm. DOM is usually quantified as “dissolved” organic carbon (DOC, mg C.L-’) since the organic matter is burned and the evolved carbon dioxide is actually quantified. In natural waters, DOM includes a wide range of constituents from truly dissolved, small molecules like acetate, to macromolecules such as humic acids, to filter- passing submicro-meter-sized particles such as viruses (Gustafsson and Gschwend, 1997).

Qualitative Description of DOM-Solute Associations

First let us consider some of the effects of DOM on the behavior of a given compound that can be observed in an aqueous solution. Very much as in the case of an organic cosolvent (Section 5.4), one can find that DOM enhances an organic compound’s apparent solubility (e.g., Chiou et al., 1986; Chin et al., 1997), and diminishes its air-water distribution ratio (Mackay et al., 1979; Brownawell, 1986). These effects are observed at DOC concentrations that are far below the concentration of any cosolvent required to produce a similar effect (e.g., 10 mg C.L-I). Furthermore, it has been demonstrated in various studies that the bioavailability of a given chemical in an aqueous solution, and thus its uptake by organisms (i.e., bioaccumulation, toxicity), may significantly decrease when DOM is added to the solution (e.g., Leversee et al., 1983; Traina et al., 1996; Haitzer et al., 1999; see also Chapter 10). Finally, it has been shown that due to interactions with DOM constituents, the light-processing characteristics of chemicals may be altered significantly. For example, it has been found that the presence of DOM diminishes the ability of polycyclic aromatic hydrocarbons (PAHs) to fluoresce (Gauthier et al., 1986 and 1987; Backhus and Gschwend, 1990). In fact, this effect has even been used in various studies for quantifying the association of PAHs with DOM.

Considering all these findings, it is reasonable to envision the interaction of organic pollutants with DOM primarily as a sorption process. That is, we consider at least some of the DOM constituents as a distinct nonaqueous organic phase. This is not surprising since filter-passing DOM certainly contains colloidal components (i.e., DOM that does not settle, but does form a “particle-like” medium). This picture is also consistent with the results of measurements of the mobility of a fluorinated probe molecule in DOM solutions using I9F-NMR (Hinedi et al. 1997). Hence, we can imagine that the most effective “DOM microsorbents” will be those that are able to accommodate a neutral organic chemical in a way that allows it to escape completely from the aqueous phase. We can look at such microsorbents as organic colloids, that is, as microparticles or macromolecules that are small enough to move primarily by Brownian motion, as opposed to gravitational settling. This means that the size of such colloids may range from about a nanometer to a micrometer in diameter. Based on this perception, the relevant DOM microsorbents in natural waters would include organic polymers as well as organic coatings on very small inorganic particles (e.g., aluminosilicates, iron oxides).

Sorption to “Dissolved” Organic Matter (DOM) 315

The K,, values for sorption of a given compound to DOM, which we will denote as KiDOC, will depend on the fraction of the DOM which is large enough to act as a sorbent (related to the molecular weight distribution of the DOM), as well as on the organic matter’s chemical structure (e.g., degree of aromaticity and presence of polar fimctionalities).

Determination of K,,, Values and Availability of Experimental Data

First, we need to comment on the available database and on the methods by which&,,, values are determined. Compared to soil-water or sediment-water systems, the determination of KLDoc values is somewhat more difficult because the organic and the aqueous phase cannot be easily separated. One separation approach uses dialysis techniques, where a solution containing the DOM is separated by a membrane from a solution containing the dissolved organic pollutant (Carter and Suffet, 1982; McCarthy and Jimenez, 1985). The membrane allows the organic compound of interest to diffuse through, but it must be impermeable for the (larger) DOM constituents. After equilibration, the partitioning substance is measured on both sides of the membrane and any “extra” concentration on the DOM-containing side is attributed to sorption to that DOM. The major drawback of this classic method ofKjDoc determination is the rather long time required to reach equilibrium between the two solutions, and the fact that some smaller DOM constituents may difise through the membrane.

Therefore, various other techniques that do not require any separation of the DOM from the aqueous phase have been applied. These methods include solubility enhancement measurements (Chiou et al., 1986; Chin et al., 1997; Hunchack- Kariouk et al., 1997), measurements of the impact of DOM on the gas exchange behavior of the solute (i.e., “gas purge techniques”; Hassett and Milicic, 1985; Luers and ten Hulscher, 1996), fluorescence-quenching techniques (Gauthier et al., 1986; Backhus and Gschwend, 1990; Schlautman and Morgan, 1993; Kumke et al., 1994; Traina et al., 1996), and direct solid-phase microextraction (SPME) of the dissolved organic species (Porschmann et al., 1997; Doll et al., 1999). All of these methods have advantages and disadvantages. The solubility enhancement method is, in principle, applicable to any kind of compound, but it yields only one point of the sorption isotherm, that is, the KiDoc value at saturation. Furthermore, the formation of microcrystals and/or emulsions may create artifacts. With the gas purge method, a large concentration range may be covered, but the method is restricted to volatile and semivolatile compounds; more importantly, due to the continuous removal of the solute, equilibrium may not always be maintained. The advantages of the fluorescence-quenching and the SPME technique are that they can cover a large concentration range. Moreover, with care they do not disturb the system significantly during the measurement. The disadvantages of the fluorescence- quenching technique are that it is restricted to fluorescent compounds and that artifacts may be produced by dynamic quenching mechanisms. Finally, the drawback of the SPME method is that it yields optimal results only for compounds that can be extracted sufficiently into the polymer film on the fiber used to concentrate the solutes from the aqueous solution.

To date, the majority of studies on the sorption of organic compounds to DOM have been conducted with PAHs. Therefore, in contrast to the situation found for soil,


phenanthrene

pyrene

aquifer, or sediment particulate organic matter, it is not possible to provide a database of KiDoc values for members of other compound classes. Nevertheless, the available data allow us to draw some important general conclusions concerning the major factors that govern sorption of apolar and weakly polar compounds to DOM. We should note that for these compounds, linear isotherms have been generally observed over the entire concentration range.

DOM Properties Governing the Magnitude of K,,,

Depending on the DOM considered, the KiDoc values of a given sorbate (e.g., PAHs like phenanthrene and pyrene) have been reported to vary by up to two orders of magnitude (Chin et al., 1997; Georgi, 1998; Laor et al., 1998; Haitzer et al., 1999; Kopinke et al., 1999). The size of the colloids or macromolecules, their conformation (coiled or extended), and their composition (e.g., aromaticity, type and number of polar groups) should influence sorption of neutral organic pollutants. Hence, in order to quantitatively describe the sorption characteristics of a given DOM, some quantitative measure of each of these bulk properties is required.

Presently the most commonly used parameters include: (1) weight-averaged or number-averaged molecular weight, (2) molar UV-light absorptivities, E~(A), at a given wavelength (e.g., il = 254 nm or 280 nm), (3) degree of aromaticity determined by 'H- and/or I3C-NMR, and (4) stoichiometric ratios of the various elements (i.e., H/C, H/O, O K , (N+O)/C) used to express the degree of saturation and the overall polarity of the DOM. Obviously, some of these parameters correlate with each other (Chin et al., 1994 and 1997; Haitzer et al., 1999). For example, the molar absorptivities at 254 or 280 nm also reflect the degree of aromaticity of a given DOM (see also Chapters 15 and 16). Furthermore, aromaticity and H/C ratio are (inversely) related. Finally, H/O, O K , and (N+O)/C all reflect the polarity of a given DOM, although they do not give any information on its H-donor and H-acceptor properties. Hence, when trying to relate KiDoc values to DOM properties, one has to be aware of possible cross-correlations. One should also be aware that these parameters represent only a crude characterization of the DOM and are, therefore, not comparable to the parameters that we have used to describe the vdW and H-bonding interactions in the polyparameter LSER models of well-defined organic so1vents:water partitioning systems (Chapter 7).

For now, we focus on bulk DOM parameters. Using pyrene as a common sorbate, 17 different DOMs were tested for their ability to sorb this PAH at 25°C. Correlations of the resultant K,,,, values with the bulk properties of the DOMs were then sought (Chin et al., 1997; Georgi, 1998). For pyrene sorption, the molar absorptivities at 280 nm (reflecting aromaticity) and the O/C ratio (reflecting overall polarity) were found to yield a significant correlation (Fig. 9.14):

log Kpyrene,DOC / (L ' kg-' OC) =

1.45 log ~ ~ ( 2 8 0 nm) / (L.mol-'oc.cm-')-1.70 (O/C)+1.14 (9-35)

( N = 17, R2 = 0.89)

It is interesting to note that the DOMs used for the derivation of Eq. 9-35 included a

Sorption to “Dissolved” Organic Matter (DOM) 317

Figure 9.14 Plot of calculated (Eq. 9-35) versus experimental K,,,, values of pyrene (pH 7, 25°C) for 17 different DOMs covering a wide range of molecular sizes, aromaticities, and polarities. Data from Chin et al. (1 997) and Georgi (1998).

3 3.5 4 4.5 5 5.5

experimental log (L-kg -’OC)

wide variety of fulvic and humic materials from creeks, rivers, swamps, lakes, soils, and dump sites, as well as commercial humic acids. In addition, 6 of the 17 Kpyrene,DOC values were determined by solubility enhancement (Chin et al., 1997), whereas the rest were measured by SPME (Georgi, 1998).

Effect of pH, Ionic Strength, and Temperature on KiDoc

pH and Ionic Strength. Because DOM constituents exhibit a rather large number of acidic (e.g., carboxylic acid, phenolic) and other polar groups that may dissociate and/or form complexes with metal ions (e.g., Na+, K’, Mg+, Ca2+, A13+), pH and major ion composition of the solution affect the sorption characteristics of a given DOM. When carboxyl groups are ionized, their electrostatic repulsions cause the DOM to spread out in solution. When divalent cations like calcium are bound to such functional groups, they enable bridging of like-charged groups and therefore cause the macromolecules to coil. The results of the limited number of studies on this subject demonstrate that these effects may be quite complex (Schlautmann and Morgan, 1993; Ragle et al., 1997; Georgi, 1998). Hence, depending on the nature of the DOM constituents, changes in pH and ion composition may or may not have a significant impact on the number of hydrophobic and hydrophilic domains that are relevant for sorption of a given neutral organic compound. Consequently, based on the very limited available data for PAHs, it is not possible to quantitatively predict the effect of solution composition on the K,,,, value of a given compound. Nevertheless, a few general comments on trends and magnitudes of such effects can be made.

First, at low to moderate ionic strengths (i.e., I < 0.1 M), we may, as a first approximation, neglect the effect of dissolved salts on KiDoc for apolar and weakly polar compounds. Since there are very little data available for higher ionic strength,


it is not possible to draw any general conclusion for seawater. We recall, however, that at high salt concentrations, salting-out phenomena as described in Chapter 5 also need to be taken into account.

When considering the effect of pH, there is a general trend that the KiDoc values of apolar and weakly polar compounds decrease with increasing pH. This effect is more pronounced for larger compounds as compared to smaller compounds. Fur- thermore, the biggest effects are found for DOMs that consist primarily of smaller sized, highly polar constituents such as fulvic acids. In these cases, differences in KiDoc of large PAHs of up to a factor of 2 to 3 or even more have been measured between pH 4 and 10 (Schlautman and Morgan, 1993; Georgi, 1998). We can rationalize these findings by envisioning an increase in the number of negatively charged functional groups with increasing pH; these may lead to the destruction of hydrophobic DOM domains (e.g., by uncoiling of macromolecules and/or by disaggrega- tion of DOM components). For the larger and less polar humic acids, however, these effects are not very substantial. Therefore, particularly within the pH-range typically encountered in natural waters (i.e., pH 6 to 9), in many cases we neglect the effect of pH On KiDOC-

Temperature. There are also very few experimental data available on the effect of temperature on DOM-water partitioning. As always, we examine the influence of temperature by considering the magnitude of the involved excess enthalpy terms in the context of an equation:

ADOM~H, 1 In KiDoc = - . - + const.

R T (9-36)

where AD OM^ Hi is the (average) standard enthalpy of transfer of i from water to the various DOM constituents and is given by:

A D O M ~ H ~ = HEOM - HI", (9-3 7)

ADoMw HI values of between -20 and -40 kJ .mol-' have been reported for six PAHs including fluoranthene and benzo(a)pyrene (Liiers and ten Hulscher, 1996). For such compounds ADOM~H, is likely dominated by the HE term. That is, we assume that HgOM is smaller since organic sorbate:organic sorbent interactions are more nearly ideal than organic so1ute:water interactions. The observed AD OM^ H, values are only slightly less negative than the corresponding negative H E values. Hence, in most cases, KID,, will be only weakly to moderately temperature dependent, that is, less than a factor of 2 with a 10 degree change in temperature. Note that a negative AD OM^ H, value means that KID,, decreases with increasing temperature.

LFERs Relating KID,,, Values to Kiow Values

Analogous to the use of simple one-parameter LFERs for relating POM-water partition coefficients (Eq. 9-25, Table 9.2) for a given DOM, we may correlate log KiDoc values with log K,,, values:

(9-38)

Figure 9.15 Equilibrium sorption to a commercial humic acid (Roth- HA): Correlations (Eq. 9-38) between log K,Doc and log K,,, for series of compounds including PAHs (open circles, log K,ooc =

0.91 log K,,, + 0.16; R2 = 0.98), some arenes including biphenyls and trans-stilbene (closed tri- angles, log K,,,, = 0.94 log K,,, - 0.29, R2 = 0.99), and alkanes (C, - CI,; open squares, log K,Do, = 0.80 K,,, + 0.24, R2 = 0.98). Note that the units of K,,,, are L.kg-' oc. Data are from Georgi (1 998).

6.0 h

8 5.5 ,-. b

c!.

g

5.0

. 4.5

k- 4.0

3.5

Y

0

m 0 -

Figure 9.16 LFERs (Eq. 9-38) relating log KjDoc and log K,,, of a set of PAHs for seven different humic or fulvic acids. Note that the units of K,,,, are L . kg-' oc. Data from Georgi (1998). The slopes and intercepts of the various LFERs are (1) 0.91/0.16, (2) 0.84/ 0.37, (3) 0.86/0.26, (4) 0.86/0.19, ( 5 ) 0.87/-0.12, (6) 0.79/0.06, (7) 0.721-0.03.

-

-

-

-

-

-

Sorption to "Dissolved" Organic Matter (DOM) 319

2 / I I I

3.0 I I I I I I I I I

3.0 3.5 4.0 4.5 5.0 5.5 6.0 6.5

Considering the complex structural features of DOM, we may expect to find good correlations only for confined sets of compounds exhibiting similar molecular interaction properties. This has been seen for the sorption of a series of PAHs, arenes, and alkanes to a commercial humic acid (Fig. 9.15). Note that linear isotherms were observed over the whole concentration range. The significant differences in KiDoc between PAHs and alkanes exhibiting the same Ki,, value may be explained by the stronger vdW interactions of the PAHs with the aromatic structures of the humic acid and/or polar interactions of the PAHs with H-donor groups.

When applying LFERs such as Eq. 9-38, the variability in compound properties and the structural variability of filter-passing natural organic matter are


important. The resultant KiDoc of a given compound may vary by more than one order of magnitude depending on the DOM considered. Hence, the slopes and intercepts of LFERs established for a given set of structurally related compounds (e.g., PAHs) for various DOMs, may be similar in some cases but may differ significantly in other cases (see examples given in Fig. 9.16). Nevertheless, because of the lack of more sophisticated approaches, in practice, simple LFERs such as Eq. 9-38 are presently the only means to make any reasonable predictions of KiDoc values. But dependable predictions (i .e~, within a factor of 2 to 3) can be expected only if an LFER is used that was established for a set of compounds and natural organic matter that is representative for the case considered (see Illustrative Example 9.5).

Illustrative Example 9.5 Evaluating the Effect of DOM on the Bioavailability of Benzo(a)pyrene (BP)

In a variety of studies, it has been shown that organic pollutants associated with DOM are not bioavailable to aquatic organisms (e.g., Leversee et al., 1993; Traina et al., 1996; Haitzer et al., 1999). Hence, the bioconcentrutionfuctor, BCFiDOC, of a given compound i in the presence of DOM can be expressed by:

where BCF, is the bioconcentration factor determined in pure water (for definitions see Chapter 10) andJ;, is the fraction that is truly dissolved in the water (see Eqs. 9-10 to 9-12).

Problem

In order to estimate how the bioavailability of benzo(a)pyrene (BP) is affected by DOM, you want to assess the speciation of this compound as a function of DOM quantity and quality. To this end, calculate the Lw value of BP for aqueous solutions (pH 7, 25°C) containing (a) 10 mg D0C.L-I and 100 mg D0C.L-', respectively, and (b) assuming DOM qualities as reflected by the LFERs #I and #7 in Fig. 9.16 (see figure caption for slopes and intercepts). Note that DOM #I represents a humic acid that exhibits a high affinity for PAHs, whereas DOM #7 is a fulvic acid with a low affinity. Hence, the two DOMs may represent extreme cases with respect to sorption of apolar and weakly polar compounds in natural waters.

Answer

When setting rsw = [DOC] and Kid = KiDOc in Eq. 9-12, the fraction in dissolved form can be calculated by:

1 1

I = benzo(a)pyrene (BP) j L W = 1 + [DOC]. K,,,,

log K,,, = 6 13 (see Appendix C)

Calculate the maximum and minimum K,,,, for BP

(1)

by inserting its K,,, value into the LFERs #1 and #7, respectively:

Sorption to Natural Organic Matter (NOM) 321

max. value: 1ogKiDoc = (0.91) (6.13) + 0.16 = 5.74

min value: log KiDoc = (0.72) (6.13) - 0.03 = 4.38

Hence, KiDoc may vary between 2.4 x 1 O4 and 5.5 x 1 O5 L . kg-’ oc. Insertion of these values into Eq. 1 (above) yieldsJw values:

for [DOC] = kg oc.L--’: between about 0.8 and 0.15

for [DOC] = lo4 kg 0c.L-’: between about 0.3 and 0.02.

This means that depending on the quantity and “quality” of the DOM present, between 2% (100 mg oc .L-’, high affinity DOM) and 80% (10 mg oc . L-’, low affinity DOM) of the total BP present may be bioavailable.

More generally, this is the fraction of BP that is available for processes that only act on the dissolved species (e.g., air-water exchange, see Chapter 20). Hence, for a strongly sorbing compound such as BP, sorption to DOM may significantly influence this PAH’s environmental behavior. In such cases, one obviously needs to get a more accurate KiDoc value for the system considered.

Sorption of Organic Acids and Bases to Natural Organic Matter (NOM)

Effect of Charged Moieties on Sorption: General Considerations

Depending on their pKia value(s), organic acids and bases may be partially or even fully ionized at ambient pH. When considering the sorption of ionized species, in addition to vdW and hydrogen bonding interactions, we also have to take into account electrostatic interactions with charged species present in natural sorbents (Fig. 9.2). These interactions are attractive if the charges of the sorbent and solute have opposite signs, and they are repulsive if both exhibit the same sign. Furthermore, some acid and base functions may also form chemical bonds with certain sorbent moieties (see example given in Fig. 9.2 and Eq. 9- 19). In these cases, quantification of sorption may become even more complicated, because the abundances and reactivities of such sorptive sites need to be quantified. We will discuss a few examples of such cases later in this section and in Chapter 1 1. Here we focus on the sorption of organic acids and bases to natural organic matter (NOM). In the following discussion we do not distinguish between sorption to DOM or POM. We quantify the overall sorption as a distribution ratio, Dioo since more than one dissolved and sorbed species may exist.

At ambient pH values, NOM is negatively charged, primarily due to the presence of carboxylic acid groups. Therefore, bulk NOM primarily acts as a cation exchanger. This means we may expect that negatively charged organic species will sorb more weakly to NOM than their neutral counterparts. Nevertheless, in situations in which


4 1 I

3 4 5 6 7 8 9 10

PH

5 1 I

G 4

b

d . 3

6' - 0 2

0 r

Y

-. 0 0

13,

t I I I 1 I

2 3 4 5 6 7 8 9

PH

Figure 9.17 Sorption of chlori- the anionic species predominates, its sorption to NOM may not always be neglected nated Phenols to sediment Organic (see below). For positively charged species, on the other hand, we may expect

stronger association with NOM constituents as compared to the neutral compound. matter: D,, values determined for (a) 2,4,5-trichlorophenoI and (b) . , pentachlorophenol- as a function ofpH. The solid line represents Eq. Sorption of Compounds Forming Anionic Species (Organic Acids) 9-41, the dashed line Eq. 9-42. Data from Schellenberg et aI. When dealing with weak acids (e.g., phenolic compounds) that exhibit only one (1984). acidic group, we may generally derive D,, using (Schellenberg et al., 1984; Lee et

al., 1990; Jahert, 1990; Severtson and Banerjee, 1996; Gundersen et al, 1997):

Since the acid and its conjugate base in solution are related via (Chapter 8):

(9-39)

(9-40)

we find:

where KEt and K z are the sorption coefficients for the neutral and the charged species, respectively, and a,, is the fraction of the compound in acidic (i~e., nondissociated) form (see Chapter 8):

1 1 + 10pH-pK'd

- - 1 a. - la - 1 + K,, / {H+}

(8-21)

As pointed out above, we may assume that Kk- will, in general, be significantly smaller than KE:. This is clearly seen for sorption of 2,4,5-trichlorophenol =

6.94) to sediment organic matter (Fig. 9 .17~) . For this type of compound, sorption of the neutral species dominates D,, up to a pH of about 2 units above the pKj, of the


PH

CI QC1

CI

2,4,5-trichlorophenoI (TCP)

PH

cl@cl CI CI

CI

pentachlorophenol (PCP)

R = C,, - C,,

&AS) linear akylbenzene sulfonates

compound. Hence, up to this pH (pK,, +2), we may neglect sorption of the anion. The overall sorption to NOM simplifies to:

Dj, = ai, . Ki"," (9-42)

Note that when applying Eqs. 9-41 and 9-42 we assume that KLA and Kkdoes not significantly depend on the pH of the solution.

At pH > - pKia + 2, particularly when dealing with hydrophobic acids such as pentachlorophenol (PCP, pKja = 4.79, the sorption of the anionic species may be important (Fig. 9.17b). In the case of PCP, the K E value is about 500 L. kg-' oc. This is somewhat less than 2 orders of magnitude smaller than the corresponding K I E value. In this case, Eq. 9-4 1 fits the experimental data well (Fig. 9.17b).

We should point out that when describing the sorption of an organic anion over a large pH-range, the pH dependence of K;",- has to be taken into account (Jafvert, 1990). This is particularly important for highly acidic, amphiphilic compounds that are present exclusively as anions at ambient pH. A prominent example involves the linear alkylbenzene sulfonates (LASs) that are widely used detergents (see also Section 2.3). For these compounds, it was found that sorption to sediments (primarily to the organic matter present) decreased by almost an order of magnitude between pH 5 and 10 (Westall et al., 1999). Furthermore, sorption isotherms were nonlinear (ni < 1 in Eq. 9-1) and were also dependent on the major cation composition of the aqueous phase. This result is not too surprising when considering that the NOM constituents (as well as other surface sites, see Chapter 11) become increasingly negatively charged with increasing pH, thus making it more difficult to accommodate a negatively charged sorbate. Similarly, we can rationalize the increasing sorption found with increasing concentration of positively charged cations (e.g., Ca"). Besides the possibility of ion pair formation with the anionic sorbate forming a neutral or even cationic species more inclined to sorb, such cations may also complex with the anionic moieties of the NOM, reducing the negative charge and the electrostatic repulsion of the anionic sorbate (Westall et al., 1999). Furthermore, nonlinear sorption may be explained by both a limited number of sites for accommodation of negatively charged species as well as increasing electrostatic repulsion with increasing sorbate concentration. Hence, all of these factors need to be considered when dealing with the sorption of anionic species. For a more detailed treatment of this topic, we refer to the literature (e.g., Jafvert, 1990; Stapleton et al., 1994; Westall et al., 1999). Finally, we should note that when dealing with sorption of surfactants such as the LASs, at higher concentrations, we also have to take into account the formation of micelles in solution as well as on surfaces (see Chapter 11).

Sorption of Compounds Forming Cationic Species (Organic Bases)

As mentioned above, sorption of the cationic form of an organic base to negatively charged sites in natural sorbents may dominate the overall sorption of the compound, at least over a certain pH-range. Examples of such cases include aliphatic and aromatic amines (Davis 1993; Lee et al., 1997; Fabrega et al., 1998), N-heterocyclic compounds (Zachara et al., 1986; Brownawell et al., 1990), and triorganotin compounds (Weidenhaupt et al., 1997; Arnold et al., 1998; Berg et al.,


Figure 9.18 Sorption of quinoline to Aldrich humic acid (AHA): Influence of pH on the overall DqDOc value (adapted from Nielsen et al., 1997).

4000

5 3000

b 0 7

Y

2000 -. 0

B 4" 1000

" 2 4 6 8 10 12

PH

2001; see also below). Because of the limited number of cation exchange sites present in a natural sorbent, sorption isotherms of the cationic species will be nonlinear and competition with other cations present (e.g., Na+, K', Mg2', Ca2', A13') may occur. Furthermore, except for the permanently negatively charged siloxane surfaces of clay minerals (see Chapter I l ) , the number of negatively charged sites present in a given natural sorbent is strongly pH-dependent. Hence, the quantitative description of sorption of organic cations to natural sorbents is a rather complex task. We will discuss some examples in more detail in Chapter 11.

Here we confine ourselves to some observations of the sorption of organic bases to NOM. Consider the pH-dependence (Fig. 9.18) of the sorption of quinoline (subscript q, for structure, see margin) to Aldrich humic acid (AHA). In this case, the DqDoc value shows a maximum at about pH 5. This corresponds to the pKi, of the compound. At high pHs (i.e., pH > 7) when virtually all of the quinoline is in its nonionic form, the overall sorption is primarily determined by partitioning of this

I H

11 pK,,= 4.90

ON> +H' .. quinoline

neutral species (Q) to AHA:

at high pHs [Qloc

[Qlw QDOC ss - = K,Q,oc (9-43)

With decreasing pH, the fraction of the cationic form of quinoline (QH') increases and the sorbed cations increase too. However, at the same time the number of negatively charged AHA moieties decreases. This leads to the maximum observed at pH 5. Now the partitioning reflects:

(9-44)

In this case, due to electrostatic interactions, the maximum D,,,, is about a factor of 4 larger than the K,QDOc (see Eq. 9-43 for partitioning of the neutral species).

An even more pronounced case involves the sorption of the two biocides, tributyltin (TBT) and triphenyltin (TPT) to AHA. Because of their very high toxicity toward


Figure 9.19 Aldrich humic acid (AHA)-water distribution ratio (Diwc) of TBT (A) and TPT (0) as a function of pH. Each point was determined from a sorption isotherm. Error bars are the standard deviation of the slope of the linear isotherm and of pH. The lines were calculated using the model described by Arnold et a]. (1998). The insert shows the speciation of TBT and TPT as a function of pH. Adapted from Arnold et al. (1997 and 1998).

11

R = n-C,H,:TBT pK,, = 6.3

R = C,H,: TPT pKi, = 5.2

5.5

h

0

7 O 5 b

=: 4.5

I! a-

Y

-I

0

0 4 0 -

3.5 ‘ I I I I I I I

3 4 5 6 7 a 9

PH

aquatic organisms, TBT and TPT are of considerable environmental concern (Fent, 1996). Again, sorption varies strongly with pH (Fig. 9.19). In these cases, the DiDoc at the pH corresponding to these compounds’ pKia values is enhanced by more than a factor of 10 over the partitioning of the neutral species (TBTOH, TPTOH). In fact, even at pH 8, where the abundance of TBT’ or TPT+ is very small, sorption of the cation was still found to dominate the overall sorption (Arnold et al., 1998).

These findings can be rationalized by postulating an inner sphere complex formation (i.e., by ligand exchange of a water molecule) between the tin atom of the charged species and negatively charged ligands (i.e., carboxylate, phenolate groups) present in the humic acid. The observed pH dependence of the overall AHA-water distribution ratio of the two compounds could be described successhlly with a semiempirical, discrete log Kj spectrum model using four discrete complexation sites in AHA exhibiting fixed pKaj values of 4,6,8, and 10 (Fig. 9.19). Note that Ki is the complexation constant of TBT’ or TPT’, respectively, with the ligand typej [i.e., a carboxyl (p& = 4,6) or phenolate (pKa, = 8,lO) group]. For more details, refer to Arnold et al. (1998).

We conclude this section by noting that sorption of charged species to NOM is generally fast and reversible, provided that no real chemical reactions take place that lead to the formation of covalent bonds (i.e., to “bound residues”; see chapter 14). This conclusion is based on experimental data and on the assumption that in aqueous solution the more polar NOM sites are more easily accessible as compared to the more hydrophobic domains. For charged species, we may, therefore, assume that equilibrium is established within relatively short time periods. Hence, for example, in the case of TBT and TPT, contaminated sediments may represent an important source for these highly toxic compounds in the overlying water column (Berg et al., 2001).


Questions and Problems

Questions

Q 9.1

Give five reasons why it is important to know to what extent a given chemical is present in the sorbed form in a natural or engineered system.

Q 9.2

What are the most important natural sorbents and sorption mechanisms for (a) apolar compounds, (b) polar compounds, and (c) ionized compounds?

Q 9.3

What is a sorption isotherm? Which types of sorption isotherms may be encountered when dealing with sorption of organic compounds to natural sorbents? Does the shape of a sorption isotherm tell you anything about the sorption mechanism(s)? If yes, what? If no, why not?

Q 9.4

Write down the most common mathematical expressions used to describe sorption isotherms. Discuss the meaning of the various parameters and describe how they can be derived from experimental data.

Q 9.5

Why is natural organic matter (NOM) such an important sorbent for all organic compounds? What types of organic phases may be present in a given system? What are the most important properties of NOM with respect to the sorption of organic compounds?

Q 9.6

How is the K,,, (K, I~oc) value of a given compound defined? How large is the variability of K,,, (K,,,,) for (a) different particulate organic phases (POM), and (b) different “dissolved” organic phases (DOM)? Which are the major structural factors of POM or DOM that cause this variability?

Q 9.7

As noted in Section 9.3 (Fig. 9 . Q the average K,,, values of 1,2-dichlorobenzene determined by Kile et al. (1 995) for uncontaminated soil-water and sediment-water partitioning are about 300 and 500 L . kg- ’ oc, respectively. However, for heavily contaminated soils and sediments, these authors found significantly higher K,,, values (700 - 3000 L.kg-’ oc), although isotherms were linear over a wide concentration range. Try to explain these findings.

Q 9.8

How do (a) pH, (b) ionic strength, and (c) temperature affect the sorption of neutral organic compounds to dissolved and particulate organic matter? Give examples of

Questions and Problems 327

compound+rganic phase combinations in which you expect (i) a minimum, and (ii) a maximum effect.

Q 9.9

How does the presence of a completely miscible organic cosolvent (CMOS) affect the speciation of an organic compound in a given environment (e.g., in an aquifer)? What are the most important parameters determining the effect of an organic cosolvent? How can this effect be quantified?

Q 9.10

What is the major difference between the sorption of neutral and the sorption of charged organic species to NOM? Qualitatively describe the pH dependence of the NOM-water partitioning of (a) an organic acid, and (b) an organic base.

Problems

P 9.1 What Fraction ofAtrazine Is Present in Dissolved Form?

Atrazine is still one of the most widely used herbicides. Estimate the fraction of total atrazine present in truly dissolved form (a) in lake water exhibiting 2 mg POC . L-’, (b) in marsh water containing 100 mg so1ids.L-’, if the solid’s organic carbon content is 20%, and (c) in an aquifer exhibiting a porosity of 0.2 by volume, a density of the minerals present of 2.5 kg.L-’, and an organic carbon content of 0.5%. Assume that partitioning to POM is the major sorption mechanism. You can find Ki,, values for atrazine in Fig. 9.9. Comment on which value(s) you select for your calculations.

CI I

H H

atrazine

P 9.2 Estimating the K,, Value of Isoproturon from Kio,’S of Structurally

Urea-based herbicides are widely used despite the concern that they may contami- nate groundwater beneath agricultural regions. You have been asked to evaluate the sorption behavior of the herbicide isoproturon in soils.

Related Compounds

I isoproturon

Unable to find information on this specific compound, you collect data on some structurally related compounds:

328 Sorption I: Sorption Processes Jnvolving Organic Matter

4-methyl 1.33 1.51

3,5 -dimethyl 1.90 1.73

4-chloro 1.94 I .95

3,4-dichloro 2.60 2.40

3-fluoro 1.37 1.73

4-methox y 0.83 1.40

What K,,, do you estimate for isoproturon? Do you use all compounds for deriving an LFER?

P 9.3 Evaluating tlze Transport oj1,2-Dichloropropane in Groundwater

A group of investigators from the USGS recently discovered a large plume of the soil fumigant 1,2-dichloropropane (DCP) in the groundwater flowing away from an airfield. The aquifer through which the DCP plume is passing has been found to have a porosity of 0.3. The aquifer solids consist of 95% quartz (density 2.65 g.mL-'; surface area 0. I m2. g '), 4% kaolinite (density 2.6 g.mL-'; surface area 10 m2. g- '), 1% iron oxides (density 3.5 g.mL-'; surface area 50 m2.g-'), and organic carbon content of 0.2%. What retardation factor [R, (= ; see Eq. 9- 12)] do you expect at mininium (assumption that only POM is responsible for sorption) for DCP transport in the plume a w n i i n g that sorptive exchanges are always at equilibrium?

1,2-dichloropropane (log K,,,=2.28. Montgomery. 1997)

P 9.4 Estimati~g the Retardation of Organic Compounds in nn Aquifer from

IJsing tritiated watei a? conservative tracer, an average retardation factor, Rf, ( ; see Eq. 9.12) of aboiit 10 was determined for chlorobenzene in an aquifer. (a) Assuming thjs retartlation factor reflects absorption only to the aquifer solids' POM, what is the avcrage organic carbon content v;,) of the aquifer material if its minerals havc density 2.5 kg.1, ' and if the porosity is 0.33? (b) Estimate the R, values of 1,3,5-tricl-iIorobei~~~~iie (1,3,5-TCB) and 2,4,6-trichlorophenol (2,4,6-TCP) in this aquifer (pFi -- 7 57 -1- - I O T ) by assuming that absorption into the POM present is the

Rreaktbrouglt Dnta of Tracer Compounds

Questions and Problems 329

major sorption mechanism. Why can you expect to make a better prediction of R, for 1,3,5-TCB as compared to 2,4,6-TCP?

You can find all necessary information in Table 9.2 and in Appendix C. Comment on all assumptions that you make.

chlorobenzene 1,3,5-trichIorobenzene 2,4,6-trichlorophenoI (1,3,5-TCB) (2,4,6-TCP)

P 9.5

Xia and Ball (1999) measured sorption isotherms for a series of chlorinated benzenes and PAHs for an aquitard material v;Oc = 0.015 kg oc . kg-' solid) from a formation believed to date to the middle to late Miocene. Hence, compared to soils or recent sediment POM, the organic matter present in this aquitard material can be assumed to be fairly mature and/or contain char particles from prehistoric fires. A nonlinear isotherm was found for TeCB (fitting Eq. 9-2) and the following Freundlich parameters were reported: KTeCB F = 128 (mg. g-')(mg. mL-')-"TecB and nTeCB = 0.80. For partitioning of TeCB to this material (linear part of the isotherm at higher concentrations), the authors found a Kim value of 4.2 x lo4 L . kg'oc.

Evaluating the Concentration Dependence of Equilibrium Sorption of 1,2,4,5-Tetrachlorobenzene (TeCB) to an Aquitard Material

(a) Calculate the apparent Ki,, values of TeCB for the aquitard material for aqueous TeCB concentrations of Ciw=l, 10, and 100 pg-L-' using the Freundlich isotherm given above. Compare these values to the K,, values given above for POM-water partitioning. Comment on the result.

(b) At what aqueous TeCB concentration (pg . L-') would the contribution of adsorption to the overall Ki,, be only half of the contribution of absorption, (partitioning)?

CI

1,2,4,5-tetrachlorobenzene (TeCB)

Note: When using Freundlich isotherms, be aware that the numerical value of KiF depends nonlinearly on the unit in which the concentration in the aqueous phase is expressed. Hence for solving this problem, you may first convert pg . L-' to mg . mL-' or you may express the Freundlich equation using, for example, pg . kg-' and pg . L-', respectively:


P 9.6 Is Sorption to Dissolved Organic Matter Important for the Environmental

Somebody claims that for naphthalene, sorption to DOM is generally unimportant in the environment. Is this statement correct? Consult Illustrative Example 9.5 to answer this question.

Behavior of Naphthalene?

naphthalene

P 9.7 Assessing the Speciation of a PCB-Congener in a Sediment-Pore Water

Consider a surface sediment exhibiting a porosity @ = 0.8, solids with average density ps = 2.0 kg . L-' solid, a particulate organic carbon content of 5%, and a DOC concentration in the pore water of 20 mg DOC. L-'. Estimate the fractions of the total 2,2',4,4'-tetrachlorobiphenyl (PCB47) present in truly dissolved form in the porewater and associated with the pore water DOM. Assume that absorption into the organic material is the major sorption mechanism and that KiDoc = 1/3 K,,,. Estimate Kioc using Eq. 9-26a with the Ki,, value of PCB47 given in Appendix C.

System

CI

2,2,4,4'-tetrachlorobiphenyl (PCB47)

Documents

Sorption I: General Introduction and Sorption Processes ...mrh/ESE 176 SGI Book/SGI Chp 9 Sorption Part 1.pdfEnvironmental Organic Chemistry, 2nd Edition. Rene P. Schwarzenbach, Philip